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WETLANDS

FIFTH EDITION

WILLIAM J. MITSCH • JAMES G. GOSSELINK

Wiley

Wetlands

Fifth Edition

William J. Mitsch
James G. Gosselink

Wiley

Dedication

This fifth edition ofMitsch and Gosselink is dedicated to my lon£i-time coauthor and friend,
Professor James G. Gosselink (1931-2015). He was ajjentleman and a scholar

whom I will^reatly miss.

— WJM

Cover image: Bernard Master
Cover design: C. Wallace

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Published by John Wiley & Sons, Inc., Hoboken, New Jersey.

Published simultaneously in Canada.

Aclaiowledgement: Figure 1.6 features four images that are and © DC Comics

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Library of Congress Cataloging-in-Publication Data:

Mitsch, William J.

Wetlands / William J. Mitsch, James G. Gosselink. — Fifth edition,
pages cm

Includes bibliographical references and index.

ISBN 978-1-118-67682-0 (cloth); 978-1-119-01978-7 (ebk.); 978-1-119-01979-4 (ebk.);

1. Wetland ecology — United States. 2. Wetlands — United States. 3. Wetland management-United
States. I. Gosselink, James G. II. Title.

QH104.M57 2015
577.68— dc23

2014040786

978-1-118-67682-0

Printed in the United States of America

10 987654321

Contents

Part I

Part II

Part III

■

Preface v

Introduction

1 Wetlands: Human Use and Science 3

2 Wetland Definitions 27

3 Wetlands of the World 45

The Wetland Environment

4 Wetland Hydrology 111

5 Wetland Soils 161

6 Wetland Blogeochemistry 179

7 Wetland Vegetation and Succession 215

Wetland Ecosystems

-m —

8 Tidal Marshes 259

9 Mangrove Swamps 311

10 Freshwater Marshes 341

11 Freshwater Swamps and Riparian Ecosystems 373

12 Peatlands 413

iv Contents

Part IV

Part V

■

Traditional Wetland Management

13 Wetland Classification 455

14 Human Impacts and Management of Wetlands 477

15 Wetland Laws and Protection 503

Ecosystem Services

16 Wetland Ecosystem Services 527

17 Wetlands and Climate Change 563

18 Wetland Creation and Restoration 591

19 Wetlands and Water Quality 647

Appendix A. Wetland Losses by State in the United
States, 1780s-1980s 701
Appendix B. Useful Wetland Web Pages 703
Appendix C. Useful Conversion Factors 705
Glossary 708
Index 721

Preface

This is the fifth edition of Wetlands — wc updated the book every seven years from
1993 to 2007 — since Van Nostrand Reinhold published the first edition in 1986.
This fifth edition (referred to here as Wetlands 5) is eight years after the Wetlands 4
but the additional one-year wait is well worth it, especially because so much new has
happened the last year in the world of wetlands.

Because of requests by many instructors using this textbook, we reincorporated
updated versions of our “ecosystem chapters” that were popular parts of the first
three editions of Wetlands. Theses ecosystem chapters — now in Part 111: Wetland
Ecosystems (chapters 8 through 12) — bring back the ecosystem view of tidal marshes,
mangroves, freshwater marshes and swamps, and northern peatlands. We had spit the
2000 edition of Wetlands 'mto essentially two books — Wetlands 4 (2007) and Wetland
Ecosystems (2009), partially because students were asking for a shorter textbook. Most
if not all of the pertinent information in those two books, all updated, is now included
in one book. Yet Wetlands Sis 744 pages long, 20 percent shorter than the 920-page
Wetlands 3. Instructors now have the choice of including or not including these
ecosystem chapters, which were always among our favorites because of their “systems”
view, in their syllabi. The chapters, by definition, integrate the otherwise separate fields
of hydrology, biogeochemistry, microbiology, vegetation, consumers, and ecosystem
function for the main types of wetlands found in the world in single chapters.

There is much new in Wetlands 5 in addition to the five reinserted and updated
ecosystem chapters in Part 111. We provide a newly published trend of wedand pub-
lications in the world, a summary and list of publications from the every-four-year
INTECOL international wetland conferences and the addition of Me^aPython vs.
Gatoroid campy science fiction wetland movie playbill to replace the long-reigning
Swamp Thin^ movie playbill in Part 1: Introduction. Updates of many of the great
wetlands of the world are also provided in this section of the book, including new
photos and descriptions of several wetlands in China.

vi Preface

Part II: The Wetland Environment (chapters 4 through 7), is significantly differ-
ent from previous editions. There are now separate new chapters, “Wetland Soils”
(Chapter 5) and “Wedand Vegetation and Succession” (Chapter 7), to complement
the updated “Wedand Hydrology” (Chapter 4) and “Wedand Biogeochemistry”
(Chapter 6) chapters. This fits better with wetland science as it is now pracdced
but also fits better the way in which we manage wedands. The book is now more
compatible with hydrology, soils, and vegetation, the three-legged stool of wetland
definitions in many countries including the United States.

The management secdon of the book is now divided into two parts: Part IV:
Tradidonal Wetland Management (chapters 13-15) and Part V: Ecosystem Services
(chapters 16-19). Chapter 13, “Wedand Classificadon,” now has an update on the
U.S. Nadonal Wedand Inventory that was just completed for the lower 48 states after
a 3 5 -year effort on May I, 2014. A web connection is also provided where readers
can obtain wedand maps from almost anywhere in the United States. A description
of methods that are being used in the United States to rate wedands is also provided
in that chapter, emphasizing systems developed in the states of Washington, Ohio,
and Florida. New peat producdon rates for countries in the world are provided in
Chapter 14, “Human Impacts and Management of Wedands,” and compared to rates
from 14 years prior. The new regional wedand delineadon manuals in the United
States are described in Chapter 15, “Wedand Laws and Protection,” as is a new U.S.
Supreme Court decision on wetland midgation that occurred in the summer of 2013.
That makes three Supreme Court decision of wetlands in the United States since the
new century began. The status of the internadonal Ramsar Convendon on Wedands,
which is growing in internadonal importance by leaps and bounds, is also brought up
to date in Chapter 1 5 .

Chapter 16, “Wedand Ecosystem Services,” now provides a new description of
the ecosystem services that wetlands provide to society, newly categorized into the
system developed by the Millennium Ecosystem Assessment of 2005 and also updates
new economic value of different types of wedands as published in mid-2014. Chapter
17, “Wetlands and Climate Change,” updates the trends in greenhouse gases in the
atmosphere and sea level rise, both of which affect and are affected by wedands. A
newly published model that provides a way of balancing the fluxes of methane with
carbon sequestradon in the same wetlands is also presented in that chapter as are
more references on these two wetland processes. We updated the “permitted” versus
“mitigated” data from the U.S. Army Corps of Engineers on trends in the USA on
midgating wedands loss in Chapter 18, “Wedand Creadon and Restoration.” We now
have seven wetland restoradon case studies thoroughly updated in this chapter: the
Florida Everglades, the Mesopotamia Marshlands in Iraq, the Bois-des-Bel experimen-
tal peadands in Quebec, Canada, the Delaware Bay and Hackensack Meadowlands salt
marsh restoradons in the eastern United States, mangrove around the Indian Ocean,
and the Skjern River channel and floodplain system in western Denmark.

Chapter 19, “Wetlands and Water Quality,” provides updates on long-term studies
that have investigated improving water quality by wedands at the Houghton Lake
treatment peadands in Michigan and the freshwater marshes at the Olentangy River

Preface vii

Wetlands in Ohio, and the wetlands filtering agricultural runoff in south Florida called
stormwater treatment areas ( STAs) . A new promising design of a stormwater treatment
wetland, located at Freedom Park in Naples Florida is presented with preliminary
data. New estimates of the costs of creating wetlands to improve water quality are also
provided in this last chapter.

We continued the tradition of “boxes” or sidebars in Wetlands 5. There are now 41
such sidebars or case studies, especially in Chapter 16, “Wetland Ecosystem Services,”
and Chapter 18, “Wetland Creation and Restoration.” New citations were added in
this edition, with over 120 from 2010 or later, to augment some classics from the past.
Many older citations, particularly those that would be hard to find, were eliminated.

On a personal note, 1 am pleased to write this edition of Wetlands from my new
venue as director and professor at the Everglades Wetiand Research Park of Florida
Gulf Coast University, located at the Naples Botanical Garden in Naples Florida.

We could not have completed this edition without help from many friends and
colleagues. Anne Mischo provided dozens more of new illustrations for Wetlands 5 to
supplement her beautiful work carried over from previous editions. We are honored
to have the cover photo of a mangrove swamp from southwest Florida, on the fringe
of the Florida Everglades, taken by a long-time friend and world-class birder Bernie
Master. Ruthmarie Mitsch provided assistance in editing some parts of this edition.
Li Zhang and Ghris Anderson need to be especially thanked for the updates that they
provided in Wetland Ecosystems that were used in this book. Li Zhang also helped on
many technical details related to publishing the book. We also appreciate the input,
illustrations, or insight provided by the following, listed in alphabetical order: Jim
Aber, Andy Baldwin, Jim Bays, Jenny Davis, Frank Day, Max Finlayson, Brij Gopal,
Glenn Guntenspergen, Wenshan He, Wolfgang Junk, David Latt, Pierrick Marion,
Mike Rochfbrd, Line Rochefort, Glayton Rubec, Kenneth Strait, and Ralph Tiner. We
also appreciate the professional effort on the part of editors and assistants at Wiley &
Sons, Inc. It has always been a pleasure to work with the Wiley brand.

Finally, you will note that there is only one author of this preface. My coauthor
Jim Gosselink, Professor Emeritus of Louisiana State University, had been ill for several
years and was unable to participate in this new edition. Jim died on January 18, 2015,
at the age of 83. But his spirit and incredible knowledge of wedands are embedded
in this book from his contributions to the previous editions, so there was no question
that his name should remain on the front of this book. I have also taken the liberty to
dedicate this book to my long-time friend and coauthor Jim Gosselink.

February 2015

William J. Mitsch, Ph.D.

Naples, Florida

Parti

Introduction

Chapter 1

Wetlands: Human Use and Science

Wetlands are found in almost all parts of the world. They are sometimes
referred to as “kidneys of the landscape” and “nature’s supermarkets” to brinp;
attention to the important ecosystem services and habitat values that they
provide. Although many cultures have lived among and even depended on
wetlands for centuries, the modern history of wetlands until the 1970s is
fraught with misunderstanding and fear, as described in much of our Western
literature. Wetlands have been destroyed at alarming rates throughout the
developed and developing worlds. Now, as their many benefits are being
recognized, wetland conservation has become the norm. In many parts of the
world, wetlands are now revered, protected, and restored; in other parts, they
are still being drained for human development.

Because wetlands have properties that are not adequately covered by
current terrestrial and aquatic ecology paradigms, a case can be made for
wetland science as a unique discipline encompassing many fields, including
terrestrial and aquatic ecology, chemistry, hydrology, and engineering. Wetland
management, as the applied side of wetland science, requires an understanding
of the scientific aspects of wetlands balanced with legal, institutional, and
economic realities. As awareness of the ecosystem services of wetlands has grown,
so too have public interest for wetland protection, wetland science programs in
universities, and publications about wetlands in scientific journals.

Wetlands are among the most important ecosystems on Earth. In the great scheme
of things, the swampy environment of the Carboniferous period produced and pre-
served many of the fossil fuels on which our society now depends. In more recent
biological and human time periods, wetlands have been valuable as sources, sinks, and
transformers of a multitude of chemical, biological, and genetic materials. Although

4 Chapter 1 Wetlands: Human Use and Science

the value of wetlands for fish and wildlife protection has been known for a century,
some of the other benefits have been identified more recendy.

Wetlands are sometimes described as kidneys of the landscape because they func-
tion as the downstream receivers of water and waste from both natural and human
sources. They stabilize water supplies, thus mitigating both floods and drought. They
have been found to cleanse polluted waters, protect shorelines, and recharge ground-
water aquifers.

Wetlands also have been called nature’s supermarkets because of the extensive
food chain and rich biodiversity that they support. They play major roles in the land-
scape by providing unique habitats for a wide variety of flora and fauna. Now that
we have become concerned about the health of our entire planet, wetlands are being
described by some as important carbon sinks and climate stabilizers on a global scale.

These values of wetlands are now recognized worldwide and have led to wetland
conservation, protection laws, regulations, and management plans. But our history
before current times with wetlands had been to drain, ditch, and fill them, never as
quickly or as effectively as was undertaken in countries such as the United States begin-
ning in the mid- 1800s. In some regions of the world that destruction of wetlands
continues.

Wetlands have become the cause celebre for conservation-minded people and
organizations throughout the world, in part because they have become symptoms of
our systematic dismantling of our water resources and in part because their disappear-
ance represents an easily recognizable loss of natural areas to economic “progress.”
Scientists, engineers, lawyers, and regulators are now finding it both useful and nec-
essary to become specialists in wetland ecology and wetland management in order
to understand, preserve, and even reconstruct these fragile ecosystems. This book is
for these aspiring wetland specialists as well as for those who would like to know
more about the structure and function of these unique ecosystems. It is a book about
wetlands — how they work and how we manage them.

Human History and Wetlands

There is no way to estimate the impact humans have had on the global extent of
wetlands except to observe that, in developed and heavily populated regions of the
world, the impact has ranged from significant to total. The importance of wetland
environments to the development and sustenance of cultures throughout human his-
tory, however, is unmistakable. Since early civilization, many cultures have learned
to live in harmony with wetlands and have benefited economically from surrounding
wetlands, whereas other cultures quickly drained the landscape. The ancient Babylo-
nians, Egyptians, and the Aztec in what is now Mexico developed specialized systems
of water delivery involving wetlands. Major cities of the world, such as Chicago and
Washington, DC, in the United States, Christchurch, New Zealand, and Paris, France,
stand on sites that were once part wetlands. Many of the large airports (in Boston,
New Orleans, and J. F. Kennedy in New York, to name a few) are situated on former
wetlands.

Human History and Wetlands 5

While global generalizations are sometimes misleading, there was and is a propen-
sity in Eastern cultures not to drain valuable wetlands entirely, as has been done in the
West, but to work within the aquatic landscape, albeit in a heavily managed way. Dugan
(1993) makes the interesting comparison between hydraulic civilizations (European
in origin) that controlled water flow through the use of dikes, dams, pumps, and
drainage tile, in part because water was only seasonally plentiful, and aquatic civiliza-
tions (Asian in origin) that better adapted to their surroundings of water- abundant
floodplains and deltas and took advantage of nature’s pulses, such as flooding. It is
because the former approach of controlling nature rather than working with it is so
dominant today that we And such high losses of wetlands worldwide.

Wetlands have been and continue to be part of many human cultures in the world.
Coles and Coles (1989) referred to the people who live in proximity to wetlands and
whose culture is linked to them as wetlanders.

Sustainable Cultures in Wetiands

Some of the original wetiander cultures are described here. The Marsh Arabs of
southern Iraq (Fig. 1.1) and the Camarguais of southern France’s Rhone River
Delta (Fig. 1.2) are two examples of ancient cultures that have lived in harmony
and sustainably with their wetland environments for centuries. In North America,
the Cajuns of Fouisiana and several Native Americans tribes have lived in harmony
with wetlands for hundreds of years. The Fouisiana Cajuns, descendants of the
French colonists of Acadia (present-day Nova Scotia, Canada), were forced out of
Nova Scotia by the English and moved to the Fouisiana delta in the last half of the

Figure 1.1 The Marsh Arabs of present-day southern Iraq lived for centuries on artificial
islands in marshes at the continence of the Tigris and Euphrates rivers in Mesopotamia. The
marshes were mostly drained by Saddam Hussein in the 1990s and are now being restored.

6 Chapter 1 Wetlands: Human Use and Science

Figure 1.2 The Camargue region of southern France in the Rhone River Deita is a historicaiiy
important wetiand region in Europe where Camarguais have iived since the Middie Ages.
(Photo by Tom Nebbia, reprinted with permission)

Figure 1.3 A Cajun iumberjack camp in the Atchafalaya Swamp of coastal Louisiana. (Photo
courtesy of the Louisiana Collection, Tkiiane University Library, reprinted with permission)

eighteenth century. Their society and culture flourished within the bayou wetlands
(Fig. 1.3). The Chippewa in Wisconsin and Minnesota have harvested and reseeded
wild rice {Zizania aquatica) along the littoral zone of lakes and streams for centuries
(Fig. 1.4). They have a saying: “Wild rice is like money in the bank.”

Likewise, several Native American tribes lived and even thrived in large-scale wet-
lands, such as the Florida Everglades. These include the ancient Calusa, a culture that
based its economy on estuarine fisheries rather than agriculture. The Calusa disap-
peared primarily as a result of imported European disease. In the nineteenth century,
the Seminoles and especially one of its tribes, the Miccosukee, moved south to the
Everglades while being pursued by the U.S. Army during the Seminole Indian wars.
They never surrendered. The Miccosukee adapted to living in hammock-style camps

Literary References to Wetlands 7

Figure 1.4 “Ricer” poling and “knocking” wiid rice [Zizania aquatica) into canoes as Anishi-
naabe (Chippewa, Ojibwe) tribes and others have done for hundreds of years on Rice Lake in
Crow Wing County, Minnesota. (Photo by John Overland, reprinted with permission)

spread throughout the Everglades and relied on fishing, hunting, and harvesting of
native fruits from the hammocks (Fig. 1.5). A recent quote in a Florida newspaper
by Miccosukee tribal member Michael Frank is poignant yet hopeful about living sus-
tainably in the Florida Everglades:

We were taught to never, ever leave the Everglades. If you leave the Everglades you
will lose your culture, you lose your language, you lose your way of life.

— Michael Frank, as quoted by William E. Gibson, “Pollution
Is Killing Everglades, Mieeosukee Warn,” South Florida
Sun Sentinel, August 10, 2013

Literary References to Wetlands

With all of these important cultures vitally depending on wedands, not to mention the
aesthetics of a landscape in which water and land often provide a striking panorama,
one might expect wetlands to be more respected by humanity; this has certainly not
always been the case. Wetlands have been depicted as sinister and forbidding and as
having little economic value throughout most of Western literature and history. For
example, in the Divine Comedy, Dante describes a marsh of the Styx in Upper Flell as
the final resting place for the wrathful:

Thus we pursued our path round a wide arc of that ghast pool.

Between the soggy marsh and arid shore.

Still eyeing those who gulp the marish [marsh] foul.

8 Chapter 1 Wetlands: Human Use and Science

Figure 1.5 The Miccesukee Native Americans adapted to iife in the Florida Evergiades in
hammock-style camps. They relied on fishing, hunting, and harvesting of native fruits from
the hammocks. (Photo by W. J. Mitsch of panorama at Miccosukee Indian Village, Florida
Evergiades)

Centuries later, Carl Linnaeus, crossing the Lapland peatlands in 1732, compared
that region to that same Styx of Hell:

Shortly afterwards began the muskegs, which mostly stood under water; these we
had to cross for miles; think with what misery, every step up to our knees. The whole
of this land of the Lapps was mostly muskeg, hinc vocavi Styx. Never can the priest
so describe hell, because it is no worse. Never have poets been able to picture Styx so
foul, since that is no fouler.

In the eighteenth century, an Englishman who surveyed the Great Dismal Swamp
on the Virginia-North Carolina border and is credited with naming it described the
wetland as

[a] horrible desert, the foul damps ascend without ceasing, corrupt the air and
render it unfit for respiration .... Never was Rum, that cordial of Life, found more
necessary than in this Dirty Place.

— Colonel William Byrd III, “Historic of the Dividing Line Betwixt
Virginia and North Carolina,” in The Westover Manuscripts,
written 1728-1736 (Petersburg, VA: E. and J. C. Ruffin, printers, 1841)

Even those who study and have been associated with wetlands have been belittled
in literature:

Hardy went down to botanise in the swamp, while Meredith climbed towards the
sun. Meredith became, at his best, a sort of daintily dressed Walt Whitman: Hardy

Literary References to Wetlands 9

became a sort of village atheist brooding and blaspheming over the
village idiot.

— G. K. Chesterton, Chapter 12 in The Victorian Age in Literature
(New York, NY: Henry Holt and Company, 1913)

The English language is filled with words that suggest negative images of wet-
lands. We get bogged down in detail; we are swamped with work. Even the mythical
bogeyman^ the character featured in stories that frighten children in many countries,
may be associated with European bogs. Grendel, the mythical monster in Beowulf^
one of the oldest surviving pieces of Old English literature and Germanic epic, comes
from the peatlands of present-day northern Europe:

Grendel, the famous stalker through waste places, who held the rolling marshes in
his sway, his fen and his stronghold. A man cut off from joy, he had ruled the
domain of his huge misshapen kind a long time, since God had condemned him in
condemning the race of Cain.

— Beowulf^ translated by William Alfred, Medieval Epics
(New York, NY: The Modern Library, 1993)

Elollywood has continued the depiction of the sinister and foreboding nature
of wedands and their inhabitants, in the tradition of Grendel, with movies such as
the classic Creature from the Black Lagoon (1954), a comic-book-turned-cult-movie
Swamp Thing {1982), and its sequel Return of the Swamp Thing {1989). Even Swamp
Thing, the man/monster depicted in Figure 1.6, evolved in the 1980s from a feared
creature to a protector of wetlands, biodiversity, and the environment. A more mod-
ern approach to scaring and entertaining the public with megafauna from the swamps
is a science fiction movie Mega Python vs. Gatoroid (2011) that is set in the Florida
Everglades (Fig. 1.7). The movie exaggerates much of the current dynamics about
the Florida Everglades including conservation, invasive species, genetically altered
organisms, fund-raising by conservationists, and conflicts among hunters, conserva-
tion agencies, and environmentalists. In some respects, current life in the Everglades
imitates art. Big snakes and alligators from wetlands continue to strike fear.

As long as wedands remain more difficult to stroll through than a forest and more
difficult to cross by boat than a lake, they will remain misunderstood by the general
public unless a continued effort of education takes place.

Food from Wetlands

Domestic wetlands such as rice paddies feed an estimated half of the world’s population
(Fig. 1.8). Gountiess other plant and animal products are harvested from wetlands
throughout the world. Many aquatic plants besides rice, such as Manchurian wild rice
{Zizania latifolia), are harvested as vegetables in Ghina. Granberries are harvested
from bogs, and the industry continues to thrive today in North America (Fig. 1.9).
Goastal marshes in northern Europe, the British Isles, and New England were used
for centuries and are still used today for grazing of animals and production of salt hay.
Salt marsh coastlines of Europe are still used for the production of salt.

10 Chapter 1 Wetlands; Human Use and Science

Figure 1.6 The sinister image of wetlands, especially swamps, has often been promoted in
popular media such as Hollywoed movies and comic books. Shown here are feur examples:
(a) Swamp Thing movie poster; (b) Swamp Thing: Dark Genesis cover; (c) Saga of Swamp
Thing #26; and (d) Swamp Thing #9. All ™ and © DC Comics.

Wetlands can be important sources of protein. The production of fish in shallow
ponds or rice paddies developed several thousands of years ago in China and Southeast
Asia, and crayfish harvesting is still practiced in the wetlands of Louisiana and the
Philippines. Shallow lakes and wedands are an important provider of protein in many
parts of sub-Saharan Africa (Fig. 1.10).

Peat and Building Materials

Russians, Finns, Estonians, and Irish, among other cultures, have mined their
peatlands for centuries, using peat as a source of energy in small-scale production

Literary References to Wetlands 11

Figure 1.7 The playbill for the Mega Python vs. Gatoroid science fiction movie pubiished by
The Asylum in 2011 (http://www.theasylum.cc). (Permissien frem David Latt, President, The
Asyium, Burbank, CA)

(Fig. 1.11) and in large-scale extraction processes (Fig. 1.12). Sphagnum peat is now
harvested for horticultural purposes throughout the world. In southwestern New
Zealand, for example, surface sphagnum has been harvested since the 1970s for
export as a potting medium. Reeds and even the mud from coastal and inland marshes
have been used for thatching for roofs in Europe, Iraq, Japan, and China as well as
in wall construction, as fence material, and for lamps and other household goods
(Fig. 1.13). Coastal mangroves are harvested for timber, food, and tannin in many
countries throughout Indo-Malaysia, East Africa, and Central and South America.

Figure 1.8 Rice production occurs in “managed” wetlands througheut Asia and other parts
of the werld. Half ef the world’s population is fed by rice paddy systems. (Photo by W. J.
Mitsch)

Figure 1.9 Cranberry wet harvesting is accemplished by flooding bogs in severai regions
of North America. The cranberry plant {Vaccinium macrocarpon) is native to the bogs and
marshes of North America and was first cultivated in Massachusetts. It is now also an impor-
tant fruit crop in Wisconsin, New Jersey, Washington, Oregon, and parts of Canada. (Photo
courtesy of Ocean Spray Cranberries, Inc.)

Literary References to Wetlands 13

Figure 1.10 Humans use the wetlands of sub-Saharan Africa for sustenance, as with this
man fishing for lungfish (Proptopterus aethiopicus) in Lake Kanyaboli, western Kenya. (Photo
by M. K. Mavuti, reprinted with permission)

Figure 1.11 Harvesting of peat, or “turf,” as a fuel has been a tradition in several parts of
the world, as shown by this scene of turf carts in Ireland.

Wetlands and Ecotourism

Eco tourism is a modem version of wetland use. Wetlands have been the focus of
attempts by several countries to increase tourist flow into their countries. The Oka-
vango Delta in Botswana is one of the natural resource jewels of Africa, and protection
of this wetland for tourists and hunters has been a priority in that country since the

14 Chapter 1 Wetlands; Human Use and Science

Figure 1.12 Large-scale peat mining in Estonia. (Phots by W. J. Mitsch)

Figure 1.13 A wetiand house in the Ebro River Delta region on the Mediterranean Sea,
Spain. The walls are made from wetland mud, and the reof is thatched with reed grass and
other wetland vegetation. (Photo by W. J. Mitsch)

1960s. Local tribes provide manpower for boat tours (in dugout canoes called moko-
ros) through the basin and assist with wildlife tours on the uplands as well (Fig. 1.14).
In Senegal, West Africa, there is keen interest in attracting European birder tourists to
the mangrove swamps along the Adantic coastline. For many people, ecotourism in
the wedands is all about the wildlife and especially the birds (Fig. 1.15). It has been
reported that bird-watching, or “birding,” is a $32 billion per year industry in the
United States alone.

Literary References to Wetlands 15

Figure 1.14 The vast seasonally flooded Okavango Delta of northern Botswana in southern
Africa is a mecca for ecotourism. The wetlands attract tourists, as shown in this illustra-
tion, and also wildlife hunting. In addition, the wetlands provides basic sustenance to these
communities. (Photo by W. J. Mitsch)

(a) (b)

Figure 1.15 Intense ecotourism interest in the wetlands in Asia is shown by (a) crowds
that surround Lake Biwa in Shiga Prefecture, Japan, at a winter 2006 international wetlands
forum, and (b) press coverage at the Ramsar Convention held in Changwon, Korea, in 2008.
(Photos by W. J. Mitsch)

The advantage of ecotourism as a management strategy is obvious — it provides
income to the country where the wedand is found without requiring or even allowing
resource harvest from the area. The potential disadvantage is that if the site becomes
too popular, human pressures will begin to deteriorate the landscape and the very
ecosystem that initially drew the tourism.

16 Chapter 1 Wetlands; Human Use and Science

Wetland Conservation

Prior to the mid-1970s, drainage and destruction of wetlands were accepted practices
around the world and were even encouraged by specific government policies. Wedands
were replaced by agricultural fields and by commercial and residential development.
Had those trends continued, wetlands would have been in danger of extinction in
some parts of the world decades ago. Some countries and states, such as New Zealand
and California and Ohio in the United States, reported 90 percent loss of their wet-
lands. Only through the combined activities of hunters and anglers, scientists and
engineers, and lawyers and conservationists has the case been made for wetlands as
a valuable resource whose destruction has serious economic as well as ecological and
aesthetic consequences for the nations of the world. This increased level of respect was
reflected in activities such as the sale of federal “duck stamps” to waterfowl hunters
that began in 1934 in the United States (Fig. 1.16); other countries, such as New

Figure 1.16 Federal Migratory Bird Hunting and Conservation Stamps are more commonly
known as duck stamps. They are produced by the U.S. Postal Service for the U.S. Fish &
Wildlife Service and are not valid for postage. Originally created in 1934 as the federal
licenses required for hunting migratory waterfowl, today income derived from their sale
is used to purchase or lease wetlands. Top; First duck stamp from 1934 (mallards); Bottom;
2013 duck stamp (wood duck).

Wetland Science and Wetland Scientists 17

Zealand, have followed suit. Approximately 2.4 million hectares (ha) ofwedands have
been purchased or leased as waterfowl habitat by the U.S. duck stamp program alone
since 1934.

The U.S. government now supports a variety of other wetland protection pro-
grams through at least a dozen federal agencies; individual states have also enacted wet-
land protection laws or have used existing statutes to preserve these valuable resources.
On an international scale, the Convention ofWedands of International Importance,
or the Ramsar Convendon, a muldnadonal agreement for the conservation of wet-
lands, has formally registered as “Wedands of Internadonal Importance” 210 million
ha of wedands in 168 contracdng parties. The Ramsar Convention is the only global
internadonal treaty specific to the conservation and wise management of a specific
ecosystem.

Wetland Science and Wetland Scientists

A specializadon in the study of wedands is often termed wetland science or wetland
ecology, and those who carry out such investigadons are called wetland scientists or
wetland ecologists. The term mire ecolo£iist has also been used. Some have suggested
that the study of all wedands be called telmatolo^y {telma being Greek for “bog”), a
term originally coined to mean “bog science” (Zobel and Masing, 1987). No matter
what the field is called, it is apparent that there are four good reasons for treadng
wedand ecology as a disdnct field of ecological study:

1 . Wedands have unique properties that are not adequately covered by present
ecological paradigms and by fields such as limnology, estuarine ecology, and
terrestrial ecology.

2. Wedand studies have begun to idendfy some common properdes of
seemingly disparate wedand types.

3. Wedand invesdgadons require a mulddisciplinary approach or training in
several fields not roudnely combined in university academic programs.

4. There is a great deal of interest in formuladng sound policy for the regulation
and management of wedands. These regulations and management approaches
need a strong sciendfic underpinning integrated as wetland ecology.

A growing body of evidence suggests that the unique characterisdcs of
wedands — standing water or waterlogged soils, anoxic conditions, and plant and
animal adaptadons — may provide some common ground for study that is neither
terrestrial ecology nor aquatic ecology. Wedands provide opportunides for tesdng
“universal” ecological theories and principles involving succession and energy flow,
theories that were developed for aquaric or terrestrial ecosystems. For example,
wedands provided the setdng for the establishment of the current system used for
lake trophic status (e.g., oligotrophic, eutrophic [Weber, 1907]), the successional
theories of Clements (1916), and the energy flow approaches of Lindeman (1942).

18 Chapter 1 Wetlands; Human Use and Science

They also provide an excellent laboratory for the study of principles related to
transition zones, ecological interfaces, and ecotones.

Our knowledge of different wetland types such as those discussed in this book is
often isolated in distinctive literatures and scientific circles. One set of literature deals
with coastal wetlands, another with forested wetlands and freshwater marshes, and
still another with peadands. Very few investigators have analyzed the properties and
functions common to all wedands. This is probably one of the most excidng areas
for wetland research because there is so much to be learned. Comparisons of wetland
types have shown, for example, the importance of hydrologic flow-through for the
maintenance and producdvity of these ecosystems. The anoxic biochemical processes
that are common to all wedands provide another area for comparadve research and
pose many quesdons: What are the roles of different wedand types in local and global
biochemical cycles.^ How do the activities of humans influence these cycles in various
wetlands.^ What are the synergisdc effects of hydrology, chemical inputs, and climatic
conditions on wetland biological productivity.^ How can plant and animal adaptadons
to anoxic stress be compared in various wetland types?

The true wetland ecologist must be an ecological generalist because of the num-
ber of sciences that bear on those ecosystems. Knowledge of wedand flora and fauna,
which are often uniquely adapted to a substrate that may vary from submerged to
dry, is necessary. Emergent wetland plant species support both aquatic animals and
terrestrial insects. Because hydrologic conditions are so important in determining
the structure and funcdon of the wedand ecosystems, a wedand scientist should be
well versed in surface and groundwater hydrology. The shallow-water environment
means that chemistry — ^pardcularly for water, sediments, soils, and water-sediment
interacdons — ^is an important science. Similarly, quesdons about wedands as sources,
sinks, or transformers of chemicals require invesdgators to be versed in many biological
and chemical techniques. While the idendficadon of wedand vegetation and animals
requires botanical and zoological skills, backgrounds in microbial biochemistry and
soil science contribute significandy to the understanding of the anoxic environment.
Understanding adaptations of wetland biota to the flooded environment requires both
biochemistry and physiology. If wedand sciendsts are to become more involved in
the management of wedands, some engineering techniques, pardcularly for wedand
hydrologic control or wedand creadon, need to be learned.

Wetlands are seldom, if ever, isolated systems. Rather, they interact strongly with
adjacent terrestrial and aquadc ecosystems. Hence, a holistic view of these complex
landscapes can be achieved only through an understanding of the principles of ecology,
especially those that are part of ecosystem and landscape ecology and systems analysis.
Finally, if wedand management involves the implementadon of wedand policy, then
training in the legal and policy-making aspects of wedands is warranted.

Thousands of scientists and engineers are now studying and managing wedands.
Only a relatively few pioneers, however, investigated these systems in any detail prior
to the 1960s. Most of the early scientific studies dealt with classical botanical sur-
veys or invesdgadons of peat structure. Several early sciendfic studies of peatland
hydrology were produced, particularly in Europe and Russia. Later, invesdgators such

Wetland Science and Wetland Scientists 19

Table 1.1 Pioneer researchers in wetiand ecology and representative citations
for their work

Wetland Type and Researcher

Country

Representative Citations

Coastal Marshes/Mangroves

Valentine J. Chapman

New Zealand

Chapman (1938, 1940)

John Henry Davis

USA

Davis (1940, 1943)

John M. Teal

USA

Teal (1958, 1962); Teal and Teal (1969)

Howard T. Odum

USA

H. T. Odum etal. (1974)

D. S. Ranwell

D. S. Ranwell (1972)

Peatlands/R’eshwater Wetlands

C. A. Weber

Germany

Weber (1907)

Herman Kurz

USA

Kurz (1928)

A. P Dachnowskl-Stokes

USA

Dachnowskl-Stokes (1935)

R. L. LIndeman

USA

Undeman (1941, 1942)

Eville Gorham

UK/ USA

Gorham (1956, 1961)

Hugo Sjors

Sweden

Sjors (1948, 1950)

G. Elnar Du Rietz

Sweden

Du Rietz (1949, 1954)

P D. Moore/D. J. Bellamy

Moore and Bellamy (1974)

S. Kuiczynski

Poland

Kuiczynski (1949)

Paul R. Errington

USA

Errington (1957)

R. S. Clymo

Clymo (1963, 1965)

Milton Weller

USA

Weller (1981)

William H. Patrick

USA

Patrick and Delaune (1972)

as Chapman, Teal, Sjors, Gorham, Eugene and H. T. Odum, Weller, Patrick, and
their colleagues and students began to use modern ecosystem and biogeochemical
approaches in wedand studies (Table 1.1). Currently active research centers devoted
to the study of wetlands include the School of Coast and Environment at Louisiana
State University; the H. T. Odum Center for Wetlands at the University of Florida;
the Duke Wetland Center at Duke University; Florida Gulf Coast University’s Ever-
glades Wetiand Research Park in Naples, Florida; the Harry Oppenheimer Okavango
Research Centre (HOORC) in Botswana, Africa: and the Institute for Land, Water,
and Society at Charles Stuart University in Australia.

In addition, a professional society now exists, the Society of Wetiand Scientists,
which has among its goals to provide a forum for the exchange of ideas within wet-
land science and to develop wetiand science as a distinct discipline. The Association
of State Wetiand Managers (ASWM) is an organization based primarily in the USA as
a place for state, federal, and local managers and consultants to meet and discuss wet-
land management issues. They currently sponsor popular webinars on subjects related
to wetlands. The International Association of Ecology (INTECOL) has sponsored
a major international wetiand conference every four years somewhere in the world
since 1980. Table 1.2 lists the locations around the world where the INTECOL Wet-
land conference has been held and each meeting’s theme, attendance, and resulting
publications.

Table 1.2 INTECOL wetland conferences, 1980 to 2012, indicating year, location, theme,
approximate attendance, chair/organizer, and resulting pubiications

Year

Location

Theme

Attendance

Organizer

Key Publication(s)

1980

New Delhi, India

B. Gopal

Gopal et al., 1982a, b

1984

Trebon,

Czechoslovakia

210

J. Kvet/J.
Pokorny

Pokorny et al., 1987;
Mitsch et al., 1988;
Bernard, 1988;
Whigham et al.,
1990, 1993

1988

Rennes, France

Conservation and
Development:
The Sustainable
Use of Wetland
Resources

400

J. C. Lefeuvre

Lefeuvre, 1989, 1990;
Maltby et al., 1992

1992

Columbus, USA

Global Wetlands:
Old World and
New

905

W. J. Mitsch

Mitsch, 1993, 1994;
Wetzel et al., 1994;
Finlayson and van der
Valk, 1995; Gopal
and Mitsch, 1995;
Jorgensen, 1995

1996

Perth, Australia

Wetlands for the
Future

550

A. J. McComb;
J. A. Davis

McComb and Davis,
1998; Tanner et al.,
1999; Zedler and
Rhea, 1998

2000

Quebec City,
Canada

Quebec 2000:
Wetlands at the
Millennium*

2160

C. Rubec, B.
Belanger,
and G. Flood

11 books/special
reports; 6 special
journal issues; 8
International Peat
Society Proceedings

2004

Utrecht,

Netherlands

787

J. T. A.
Verhoeven

Vymazal, 2005;

Bobbink et al., 2006;
Junk 2006; van
Diggelen et al., 2006;
Verhoeven et al.,
2006; Davidson and
Finlayson, 2007;
Whitehouse and
Bunting, 2008

2008

Cuiaba, Brazil

Big Wetlands, Big
Concerns

700

P Teixeira de
Sousa Jr.; C.
Nunes da
Cunha

Vymazal, 2011; Junk,
2013

2012

Orlando, USA

Wetlands in a
Complex World**

1240

R. Best/ K.R.
Reddy

•INTECOL met with three additional societies in 2000: Internationai Peat Society; International Mire Conserva-
tion Group; Society of Wetland Scientists.

••iNTECOL met with Society of Wetiand Scientists in 2012.

Wetland Managers and Wetland Management 21

Year

Figure 1.17 Science Citatien Index (SCI) listed scientific articles that included “wetland” in
their title or in keywords during the period 1960 to 2010 (From Zhang et al., 2010).

The increasing interest and emphasis on wetland science and management has
been demonstrated by a veritable flood of books, reports, scientiflc journal articles,
and conference proceedings, most in the last two decades of the twentieth century
and the first decade of the twenty-first century. From 1991 to 2008, the annual num-
ber of wetland research journal articles published and the number of wetland articles
cited increased six- and nine-fold, respectively (Fig. 1.17). The journal citations in this
book are only the tip of the iceberg of the literature on wetlands. Two journals specific
to wetlands — Wetlands Wetlands Ecology and Mana£fement — are now published to
disseminate scientiflc and management papers on wetlands, and several other scholarly
journals frequendy publish papers on the topic. Dozens of wetland meeting proceed-
ings and journal special issues have been published from conferences on wetlands held
throughout the world.

Wetland Managers and Wetland Management

Just as there are wedand sciendsts who are uncovering the processes that determine
wedand funcdons and values, so too there are those who are involved, by choice or
by vocadon, in some of the many aspects of wedand management. These individuals,
whom we call wetland mana^ers^ are engaged in acdvides that range from waterfowl
producdon to wastewater treatment. They must be able to balance the sciendfic aspects

22 Chapter 1 Wetlands; Human Use and Science

of wetlands with myriad legal, institutional, and economic constraints to provide opti-
mum wetland management. The management of wedands has become increasingly
important in many countries because government policy and wetland regulation seek
to reverse historic wetland losses in the face of continuing draining or encroachment
by agricultural enterprises and urban expansion. The simple act of being able to iden-
tify the boundaries of wetlands has become an important skill for a new type of wetland
technician in the United States called a wetland delineator.

Private organizations, such as Ducks Unlimited, Inc. and The Nature Con-
servancy have protected wetlands by purchasing thousands of hectares of wedands
throughout North America. Through the Ramsar Convendon and an agreement
joindy signed by the United States and Canada in 1986 called the North American
Waterfowl Management Plan, wedands are now being protected primarily for their
waterfowl value on an internadonal scale. In 1988, a federally sponsored Nadonal
Wedands Policy Forum (1988) in the United States raised public and polidcal
awareness of wedand loss and recommended a policy of “no net loss” of wedands.
This recommendadon has sdmulated widespread interest in wedand restoradon
and Creadon to replace lost wetlands, and “no net loss” has remained the policy of
wetland protection in the United States since the late 1980s.

Subsequendy, a Nadonal Research Council (NRC) report in the United States
(NRC, 1992) called for the fulfillment of an ambidous goal of gaining 4 million ha of
wetlands by the year 2010, largely through the reconversion of crop- and pastureland.
That goal was not met. Wedand creadon for specific funcdons remains an excidng
new area of wedand management that needs trained specialists and may eventually
stem the tide of loss and lead to an increase in this important resource. Another NRC
report (1995) reviewed the scientific basis for wedand delineadon and classification,
particularly as it related to the reguladon of wetlands in the United States at that dme,
and yet another NRC (200 1 ) study invesdgated the effectiveness of the nadonal policy
of mitigation of wedand loss in the United States.

Wetland management organizadons, such as the Associadon of State Wetland
Managers (ASWM) and the Society of Wedand Sciendsts (SWS), focus on dissem-
inadng informadon on wedands, pardcularly in North America. The Internadonal
Union for the Conservation of Nature and Natural Resources (lUCN) and the Ram-
sar Convendon, both based in Switzerland, have developed a series of publicadons
on wetlands of the world. Wetlands International (www.wetlands.org) is the world’s
leading nonprofit organizadon concerned with the conservadon of wedands and wet-
land species. It comprises a global network of governmental and nongovernmental
experts working on wedands. Acdvities are undertaken in more than 120 countries
worldwide. The head office is located in Wageningen, Netherlands.

Recommended Readings

Beaudfully illustrated popular books and articles, many with color photographs, were
developed on wetlands by many authors in years past. Here are some of our dmeless
favorites.

Dugan, R 1993. Wetlands in Danger. London: Oxford University Press.

Finlayson, M., and M. Moser, eds. \991. Wetlands. Oxford, UK: Facts On File.

References 23

Kusler, J., W. J. Mitsch, and J. S. Larson. 1994. Wetlands. Scientific American 270(1):
64-70.

Litdehales, B., and W. A. Niering. 1991. Wetlands ofi North America. Charlottesville,
VA: Thomasson-Grant.

Lockwood, C. C., and R. Gary. 2005. Marsh Mission: Capturing the Vanishing Wet-
lands. Baton Rouge: Louisiana State University Press.

McGomb, A. J., and P. S. Lake. 1990. Australian Wetlands. London: Angus and
Robertson.

Mendelsohn, J., and S. el Obeid. 2004. Okavango River: The Flow of a Lifeline. Gape
Town, South Africa: Struik.

Mitchell, J. G., R. Gehman, and J. Richardson. 1992. Our Disappearing Wetlands.
National Geographic 182(4): 3-45.

Niering, W. A. 1985. Wetlands. New York: Knopf

Rezendes, P, and P. Roy. 1996. Wetlands: The Web of Life. San Francisco: Sierra Glub
Books.

References

Bernard, J. M., ed. 1998, Carex. Special Issue of Aquatic Botany 30: 1-168.

Bobbink R., B. Beltman, J. T. A. Verhoeven, and D. F. Whigham, eds. 2006. Wedands:
Functioning, Biodiversity, Gonservation and Restoration. Ecolopfical Studies 191,
Springer, Berlin, 315 pp.

Ghapman, V. J. 1938. Studies in salt marsh ecology. ITll. Journal of Ecology 26:
144-221.

Ghapman, V. J. 1940. Studies in salt marsh ecology. Vl-Vll. Journal of Ecology 28:
118-179.

Glements, F. E. 1916. Plant Succession. Publication 242. Garnegie Institution of
Washington. 512 pp.

Glymo, R. S. 1963. Ion exchange in Sphapfnum and its relation to bog ecology. Annals
of Botany (London) New Series 27: 309-324.

Glymo, R. S. 1965. Experiments on breakdown of Sphaanum in two bogs. Journal of
Ecology 53: 747-758.

Goles, B., and J. Goles. 1989. People of the Wetlands, Bojjs, Bodies and Lake-Dwellers.
Thames Hudson, New York. 215 pp.

Dachnowski-Stokes, A. P. 1935. Peat land as a conserver of rainfall and water supplies.
Ecolosy 16: 173-177.

Davidson, N. G., and M. Finlayson, eds. 2007. Satellite-based radar - Developing
tools for wetlands management. Aquatic Conservation: Marine and Freshwater
Ecosystems 17(3): 219-329.

Davis, J. H. 1940. The ecology and geologic role of mangroves in Florida. Publication
517. Garnegie Institution of Washington, pp. 303-412.

Davis, J. H. 1943. The natural features of southern Florida, especially the vegetation
and the Everglades. Elorida Geolopfical Survey Bulletin 25. 311 pp.

Dugan, P. 1993. Wetlands in Danpier. Michael Beasley, Reed International Books,
London. 192 pp.

24 Chapter 1 Wetlands; Human Use and Science

Du Rietz, G. E. 1949. Huvudenheter och huvudgranser i Svensk myrvegetation.
Svensk Botanisk Tidkrift 43'. 274-309.

Du Rietz, G. E. 1954. Die Mineralbodenwasserzeigergrenze als Grundlage Einer
Natiirlichen Zweigleiderung der Nord-und Mitteleuropaischen Moore. Vejjetatio
5-6: 571-585.

Errington, R L. 1957. Of Men and Marshes. The Iowa State University Press,
Ames, lA.

Finlayson, G. M. and A. G. van der Valk. 1995. Glassification and inventory of the
world’s wetlands. Special Issue Ve^etatio 118: 1-192.

Gopal, B., R. E. Turner, R. G. Wetzel, and D. F. Whigham, eds. 1982a. Wedands:
Ecology and Management. International Scientific Publications, Jaipur, India.
Vol. 1, 514 pp.

Gopal, B., R. E. Turner, R. G. Wetzel, and D. F. Whigham, eds. 1982b. Wetlands:
Ecology and Management. International Scientific Publications, Jaipur, India.
Vol. 2, 156 pp.

Gopal, B., and W. J. Mitsch, eds. 1995. The role of vegetation in created and restored
wetlands. Special Issue Ecological En^ineerinp; 5 : 1-12 1 .

Gorham, E. 1956. The ionic composition of some bogs and fen waters in the English
lake district. Journal of Ecology 44: 142-152.

Gorham, E. 1961. Factors influencing supply of major ions to inland waters, with
special references to the atmosphere. Geological Society of America Bulletin 72:
795-840.

Jorgensen, S. E., ed. 1995. Wedands: Interactions with watersheds, lakes, and riparian
zones. Special issue Wetlands Ecology and Manap/ement 3:79-137.

Junk, W., ed. 2006. The comparadve biodiversity of seven globally important
wedands. Special Issue of Aquatic Sciences 68(3): 239M:14.

Junk, W., ed. 2013. The world’s wetlands and their future under global climate
change. Special Issue of Aquatic Sciences 75(1): 1-167.

Kulczynski, S. 1949. Peat bogs of Polesie. Acad. Pol. Sci. Mem.., Ser. B, No. 15.
356 pp.

Kurz, H. 1928. Influence of Sphagnum and other mosses on bog reactions. Ecology
9: 56-69.

Lefeuvre J. G., ed .1989. Gonservadon et developpement : gesdon integree des zones
humides. Troisieme conference internadonale sur les zones humides, Rennes,
19-23 Septembre 1988. Ed. Museum Nadonal d’Histoire Naturelle, Laboratoire
d’Evoludon des Systemes Naturels et Modifies, Paris. 371 pp.

Lefeuvre J. G. 1990. INTEGOL’s Third Internadonal Wetlands Gonference. Rennes,
1988. Bull.EcoL, 21(3), 80 pp.

Lindeman, R. L. 1941. The developmental history of Gedar Greek Lake, Minnesota.
American Midland Naturalist 25: 101-112.

Lindeman, R. L. 1942. The trophic -dynamic aspect of ecology. Ecology 23: 399-418.

Maltby E., P. J. Dugan, and J. G. Lefeuvre, eds. 1992. Gonservadon and Devel-
opment: The Sustainable Use of Wedand Resources. Proceedings of the Third
Internadonal Wedands Gonference. lUGN, Gland, Switzerland. 219 pp.

References 25

McComb, A. J. and J. A. Davis, eds. 1998. Wetlands for the Future - Contribu-
tions from INTECOL’s V International Wetlands Conference. Gleneagles Press,
Adelaide, Australia, 750 pp.

Mitsch, W. J. 1993. INTECOL’s IV International Wetlands Conference: A report.
International Journal of Ecology and Environmental Sciences 19:129-134.

Mitsch, W. J., ed. 1994. Global Wetlands: Old World and New. Elsevier, Amsterdam.
967+ xxiv pp.

Mitsch, W. J., and J. G. Gosselink. 1986. Wetlands, Van Nostrand Reinhold, New
York. 539 pp.

Mitsch, W. J., M. Straskraba, and S.E. Jorgensen, eds. 1988. Wedand Modelling.
Elsevier, Amsterdam, 227 pp.

Mitsch, W. J., and J. G. Gosselink. 1993. Wetlands, 2nd ed. Van Nostrand Reinhold
and John Wiley 8e Sons, New York. 722 pp.

Mitsch, W. J., and J. G. Gosselink. 2000. Wetlands, 3rd ed. John Wiley & Sons, New
York. 920 pp.

Mitsch, W. J., and J. G. Gosselink. 2007. Wetlands, 4th ed. John Wiley 8e Sons, Hobo-
ken, NJ. 582 pp.

Moore, P. D., and D. J. Bellamy. 1974. Peatlands. Springer-Verlag, New York. 221 pp.

National Research Gouncil (NRG). 1992. Restoration of Aquatic Ecosystems . National
Academy Press, Washington, DG. 552 pp.

National Research Gouncil (NRG). 1995. Wetlands: Characteristics and Boundaries.
National Academy Press, Washington, DG. 306 pp.

National Research Gouncil (NRG). 2001. Compensating for Wetland Losses under the
Clean Water Act . National Academy Press, Washington, DG, 158 pp.

National Wetlands Policy Forum. 1988. Protecting America’s Wetlands: An Action
Apienda. Gonservation Foundation, Washington, DG. 69 pp.

Odum, H. T, B. J. Gopeland, and E. A. McMahan, eds. 1974. Coastal Ecological
Systems of the United States. Gonservation Foundation, Washington, DG. 4 vols.

Patrick, W. H., Jr., and R. D. Delaune. 1972. Gharacterization of the oxidized and
reduced zones in flooded soil. Proceedinpis of the Soil Science Society of America 36:
573-576.

Pokorny, J., O. Lhotsky, P. Denny, and R. E. Turner, eds. 1987. Waterplants and
wedand processes. Special issue of Archiv Pur Hydrohiologie 27 : 1-VIII and 1-265 .

Ranwell, D. S. 1972. Ecolopsy of Salt Marshes and Sand Dunes. Ghapman & Hall,
London. 258 pp.

Sjors, H. 1948. Myrvegetation i bergslagen. Acta Phyto^eo^raphica SuecicalU. 1-299.

Sjors, H. 1950. On the relationship between vegetation and electrolytes in North
Swedish mire waters. Oikos 2: 239-258.

Tanner, G. G., G. Raisin, G. Ho, and W. J. Mitsch, eds. 1999. Gonstructed and Natural
Wetlands for Pollution Gontrol. Special Issue of Ecological Enpfineerinpi 12: 1-170.

Teal, J. M. 1958. Distribution of fiddler crabs in Georgia salt marshes. Ecology 39:
18-19.

Teal, J. M. 1962. Energy flow in the salt marsh ecosystem of Georgia. Ecology 43:
614-624.

26 Chapter 1 Wetlands; Human Use and Science

Teal, J. M., and M. Teal. 1969. Life and Death of the Salt Marsh. Little, Brown, Boston.
278 pp.

van Diggelen, R., Middleton, B., Bakker, J.R, Grootjans, A.R, Wassen, M.J. (eds.)
2006. Fens and floodplains of the temperate zone: Rresent status, threats, conser-
vation and restoration. Special Issue of Applied Vegetation Science 9(2): 157-316.

Verhoeven J. T. A., B. Beltman, R. Bobbink, and D. F. Whigham, eds. 2006. Wedands
and Natural Resource Management. Ecological Studies 190, Springer, Berlin,
347 pp.

Vymazal, J., ed. 2005. Constructed wetlands for wastewater treatment. Special Issue
of Ecolojjical Enjjineerinpf 25: 475-621.

Vymazal, J., ed. 2011. Enhancing ecosystem services on the landscape with created,
constructed and restored wetlands. Special Issue of Ecoloptical Enptineerinpt 37(1):
1-98.

Weber, C. A. 1907. Autbau und Vegetation der Moore Norddutschlands. Beibl. Bot.
Jahrb. 90: 19-34.

Weller, M. W. 1981. Freshwater Marshes . University of Minnesota Rress, Minneapolis.
146 pp.

Wetzel, R. G., A. van der Valk, R. E. Turner, W. J. Mitsch and B. Gopal, eds. 1994.
Recent studies on ecology and management of wetlands. Special Issue of Interna-
tional Journal of Ecology and Environmental Sciences 20(1-2): 1-254.

Whigham, D. F., R. E. Good, and J. Kvet, eds. 1990. Wetland Ecology and Manage-
ment: Case Studies, Tasks for Vegetation Science 23, Kluwer Academic Rublishers,
Dordrecht, 180 pp.

Whigham, D. F., D. Dykyjova, and S. Hejny, eds. 1993. Wedands of the World:
Inventory, Ecology and Management. Volume 1. Africa, Australia, Canada and
Greenland, Mediterranean, Mexico, Rapua New Guinea, South Asia, Tropical
South America, United States. Kluwer Academic Rublishers, Dordrecht, 768 pp.

Whitehouse, N. J., and M.J. Bunting, eds. 2008. Ralaeoecology and long-term wet-
land function dynamics: A tool for wetland conservation and management. Special
Issue of Biodiversity and Conservation 17(9): 2051-2304.

Zedler, J., and N. Rhea, eds. 1998. Ecology and management of wetland plant inva-
sions. Special issue of Wetlands Ecolopiy and Management 5(3): 159-242.

Zhang. L., M. H. Wang, J. Hu, and Y.-S. Ho. 2010. A review of published wetland
research, 1991-2008: Ecological engineering and ecosystem restoration. Ecolog-
ical Engineering 36: 973-980.

Zobel, M., and V. Masing. 1987. Bog changing in time and space. Archiv fur Hydro-
biologie, Beiheft: Ergebnisse der Limnologie 27: 41-55.

Chapter 2

Wetland Definitions

Wetlands have many distin£imshin£i features, the most notable of which are the
presence of standing water for some period durinp; the pirowinpi season, unique
soil conditions, and organisms, especially vegetation, adapted to or tolerant of
saturated soils. Wetlands are unique because of their hydrologic conditions and
their role as ecotones between terrestrial and aquatic systems. Terms such as
swamp, marsh, fen, and bog have been used in common speech for centuries to
define wetlands and are frequently used and misused today. Formal definitions
have been developed by scientists and federal agencies in the United States and
Canada and through an international treaty known as the Ramsar
Convention. These definitions are used for both scientific and management
purposes. Wetlands are not easily defined, however, especially for legal purposes,
because they have a considerable range of hydrologic conditions, because they are
found along a gradient at the margins of well-defined uplands and deepwater
systems, and because of their great variation in size, location, and human
influence. No absolute answer to “What is a wetlandl” should be expected, but
legal definitions involving wetland protection have become the norm.

The most common questions that the uninitiated ask about wetlands are “What exactiy
is a wetiand?” or “Is that the same as a swamp?” These are surprisingly good questions,
and it is not altogether clear that they have been answered completely by wetland sci-
entists and managers. Wetland definitions and terms are many and are often confusing
or even contradictory. Nevertheless, definitions are important both for the scientific
understanding of these systems and for their proper management.

In the nineteenth century, when the drainage of wedands was the norm, a wetland
definition was unimportant because it was considered desirable to produce uplands
from wetlands by draining them. In fact, the word wetland did not come into common
use until the mid-twentieth century. One of the first references to the word was in the

28 Chapter 2 Wetland Definitions

publication Wetlands of the United States (Shaw and Fredine, 1956). Before that time,
wetlands were referred to by the many common terms that developed in the nineteenth
century and before, such as swamp, marsh, bop;, fen, mire, and moor. Even as the value
of wedands was being recognized in the early 1970s, there was little interest in precise
definitions until it was realized that a better accounting of the remaining wetland
resources was needed, and definitions were necessary to achieve that inventory.

When national and international laws and regulations pertaining to wetland
preservation began to be written in the late 1970s, the need for precision became
even greater as individuals recognized that definitions were having an impact on
what they could or could not do with their land. The definition of a wetland, and
by implication its boundaries (referred to as “delineation” in the United States),
became important when society began to recognize the value of these systems and
began to translate that recognition into laws to protect itself from further wetland
loss. However, just as an estimate of the boundary of a forest, desert, or grassland is
based on scientifically defensible criteria, so too should the definition of wetlands be
based on scientific measures to as great a degree as possible. What society chooses to
do with wetlands, once the definition has been chosen, remains a political decision.

Wetlands in the Landscape

Even after the ecological and economic benefits of wetlands were determined and
became widely appreciated, wetlands have remained an enigma to scientists. They are
difficult to define precisely, not only because of their great geographical extent but
also because of the wide variety of hydrologic conditions in which they are found.
Wetlands are usually found at the interface of terrestrial ecosystems, such as upland
forests and grasslands, and aquatic systems, such as deep lakes and oceans (Fig. 2.1a),
making them different from each yet highly dependent on both. They are also found in
seemingly isolated situations, where the nearby aquatic system is often a groundwater
aquifer (Fig. 2.1b). Sometimes these wedands are referred to as isolated wetlands,
a somewhat misleading term because they are usually connected hydrologically to
groundwater and biologically through the movement of many mobile organisms. And,
of course, all wedand ecosystems are open to solar radiadon and precipitation.

Because wedands combine attributes of both aquadc and terrestrial ecosystems
but are neither, they have fallen between the cracks of the scientific disciplines of
terrestrial and aquatic ecology. They serve as sources, sinks, and transformers of nutri-
ents; deepwater aquatic systems (at least lakes and oceans) are almost always sinks, and
terrestrial systems are usually sources. Wedands are also among the most producdve
ecosystems on the planet when compared to adjacent terrestrial and deepwater aquadc
systems, but it is not correct to say that all wedands are highly producdve. Peadands
and cypress swamps are examples of low-producdvity wetlands.

Distinguishing Features of Wetiands

We can easily idendfy a coastal salt marsh, with its great uniformity of grasses and its
maze of ddal creeks, as a wetland. A cypress swamp, with majesdc trees festooned

Wetlands in the Landscape 29

with Spanish moss and standing in knee-deep water, provides an unmistakable image
of a wetland. A northern Sphagnum bog, surrounded by tamarack trees that quake as
people trudge by, is another easily recognized wetland. All of those sites have several
features in common: (1) all have shallow water or saturated soil; (2) all accumulate
organic plant material that decomposes slowly; and (3) all support a variety of plants
and animals adapted to the saturated conditions. Wetland dehnitions, then, often
include three main components:

1 . Wedands are distinguished by the presence of water, either at the surface or
within the root zone.

2. Wetlands often have unique soil conditions that differ from adjacent uplands.

3. Wetlands support biota such as vegetation adapted to the wet conditions
{hydrophytes) and, conversely, are characterized by an absence of
flooding-intolerant biota.

TERRESTRIAL

SYSTEM

WETLAND

DEEPWATER
AQUATIC SYSTEM

Figure 2.1 Wetlands are often located (a) between dry terrestrial systems and permanently
flooded deepwater aquatic systems such as rivers, lakes, estuaries, or oceans or (b) as iso-
lated basins with little outflow and no adjacent deepwater system.

30 Chapter 2 Wetland Definitions

TERRESTRIAL WETLAND

SYSTEM

Figure 2.1 {Continued)

This three-level approach to the definition of wetlands is illustrated in Figure 2.2.
Climate and geomorphology define the degree to which wedands can exist, but the
starting point is the hydrolqgiy, which, in turn, affects the physiochemical environment,
including the soils, which, in turn, determines with the hydrology what and how much
biota, including vegetation, is found in the wetland. This model is reintroduced and
discussed in more detail in Chapter 4: “Wetland Hydrology.”

Difficulty of Defining Wetlands

Although the concepts of shallow water or saturated conditions, unique wetland soils,
and vegetation adapted to wet conditions are fairly straightforward, combining these

Wetlands in the Landscape 31

Figure 2.2 The three-component basis of a wetiand definition: hydrology, physiochemical
environment, and biota. Ftom these compenents, the current approach to defining juris-
dictional wetlands in the United States is based en three indicators: hydrology, soils, and
vegetation. Note that these three components are not independent and that there is signifi-
cant feedback frem the biota to the physics, chemistry, and hydrology.

three factors to obtain a precise definition is difficult because of six characteristics that
distinguish wedands from other ecosystems yet make them less easy to define:

1. Although water is present for at least part of the time, the depth and duration of
flooding vary considerably from wetland to wetland and from year to year.

Some wetlands are continually flooded, whereas others are flooded only
briefly at the surface or even just below the surface. Similarly, because
fluctuating water levels can vary from season to season and year to year in the
same wedand type, the boundaries of wetlands cannot always be determined
by the presence of water at any one dme.

2. Wetlands are often located at the marpfins beWeen deep water and terrestrial
uplands and are influenced by both systems. This ecotone posidon has been
suggested by some as evidence that wedands are mere extensions of either
the terrestrial or the aquadc ecosystems or both and have no separate
idendty. There are, however, emergent properdes in wetlands not contained
in either upland or deepwater systems.

32 Chapter 2 Wetland Definitions

3. Wetland species (plants, animals, and microbes) ran^efrom those that have
adapted to live in either wet or dry conditions (facultative), which makes
difficult their use as wetland indicators, to those adapted to only a wet
environment (oblipfate).

4. Wetlands vary widely in size, ranpfin^ from small prairie potholes of a few
hectares in size to larpie expanses of wetlands several hundreds of square
kilometers in area. Although this range in scale is not unique to wetlands, the
question of scale is important for their conservation. Wetlands can be lost in
large parcels or, more commonly, one small piece at a time in a process called
cumulative loss. Are wedands better defined functionally on a large scale or in
small parcels.^

5. Wetland location can vary greatly, from inland to coastal wetlands and from
rural to urban regions. Whereas most ecosystem types — for example, forests
or lakes — have similar ecosystem structure and function, there are great
differences among different wetland types such as coastal salt marshes, inland
pothole marshes, and forested bottomland hardwoods.

6. Wetland condition, or the decree to which a wetland has been modified by
humans, variesptreatlyfrom region to region and from wetland to wetland. In
rural areas, wetlands are likely to be associated with farmlands, whereas
wetlands in urban areas are often subjected to the impact of extreme pollution
and altered hydrology associated with housing, feeding, and transporting a
large population. Many wetlands can easily be drained and turned into dry
lands by human intervention; similarly, altered hydrology or increased runoff
can cause wetlands to develop where they were not found before.

Wetlands have been described as a halfway world between terrestrial and aquatic
ecosystems, exhibiting some of the characteristics of each system. They form part of a
continuous gradient between uplands and open water. As a result, the exact upper and
the lower limits of wetlands are arbitrary boundaries in any definition. Consequently,
few definitions adequately describe all wetlands.

The problem of definition arises at the edges of wetlands, toward either wetter or
drier conditions. How far upland and how infrequently should the land flood before
we can declare that it is not a wetland.^ At the other edge, how far can we venture into
a lake, pond, estuary, or ocean before we leave a wetland.^ Does a floating mat of veg-
etation define a wetland.^ What about a submerged bed of rooted vascular vegetation.^

The frequency of flooding is another variable that has made the definition of
wetlands particularly controversial. Some classifications include seasonally flooded bot-
tomland hardwood forests, whereas others exclude them because they are dry for most
of the year. Because wetland characteristics grade continuously from aquatic to terres-
trial, there is no single, universally recognized definition of a wetland. This lack has
caused confusion and inconsistencies in the management, classification, and invento-
rying of wetland systems, but considering the diversity of types, sizes, locations, and
conditions of wetlands in this country, inconsistencies should be no surprise.

Wetland Common Terms 33

Wetland Common Terms

A number of common terms have been used over the years to describe different types
of wetlands (Table 2.1). The number of common wedand words has risen from the
15 listed in the first edition of Wetlands (Mitsch and Gosselink, 1986) to 40 terms
in this edition as we continue to discover terms in use. The history of the use and
misuse of these words has often revealed a decidedly regional or at least continental
origin. Although the lack of standardization of terms is confusing, many of the old
terms are rich in meaning to those familiar with them. They often bring to mind vivid
images of specific kinds of ecosystems that have distinct vegetation, animals, and other
characteristics. Each of the terms has a specific meaning to some people, and many are
still widely used by both scientists and laypersons alike. A marsh is known by most as an
herbaceous plant wetland. A swamps however, has woody vegetation, either shrubs or
trees. There are subtle differences among marshes. A marsh with significant (>30 cm)
standing water throughout much of the year is often called a deepwater marsh. A
shallow marsh with waterlogged soil or shallow standing water is sometimes referred
to as a sed^e meadow or a wet meadow. Intermediate between a marsh and a meadow is
a wet prairie. Several terms are used to denote peat-accumulating systems. The most
general term is peatland., which is generally synonymous with moor and muskeg. There
are many types of peatlands, however, the most general being /ewr and bojjs.

Within the international scientific community, these common terms do not always
convey the same meaning relative to a specific type of wetland. In fact, some languages
have no direct equivalents for certain kinds of wedands. The word sivamp has no
direct equivalent in Russian because the forested wedands there are simply a variety of
peadands or bogs. Bojj, however, can easily be translated because bogs are a common
feature of the Russian landscape. The word swamp in North America clearly refers
to a wedand dominated by woody plants — shrubs or trees. In Europe, reedswamps
are dominated by reed grass {Phragmites)^ a dense-growing but nonwoody plant. In
Africa, what would be called a marsh in the United States is referred to as a swamp. A
cutoff meander of a river is called a hillabong 'm. Australia (Shiel, 1994) and an oxbow
in North America.

Even common and scientific names of plants and animals can become confusing on
a global scale. Typha spp., a cosmopolitan wedand plant, is called cattail in the United
States, reedmace in the United Kingdom, bulrush in Africa, eumhungi in Australia,
and raupo or bulrush in New Zealand. True bulrush is sdll called Scirpus spp . by some
in North America and Schoenoplectus spp. in much of the rest of the world. Scirpus
fluviatilis (river bulrush) is Bolboschoenus fluviatilis in much of the rest of the world.
The great egret in North America is Casmerodius albus., whereas the great egret in
Australia is Ardea alba. To further complicate matters, the Australian version of the
great egret is called Epretta alba in New Zealand and is not called an egret at all but
a white heron.

Confusion in terminology occurs because of different regional or condnental uses
of terms for similar types of wedands. In North America, nonfbrested inland wet-
lands are often casually classified either as peat-forming, low-nutrient acid bogs or as

Table 2.1 Common terms used for various wetland types in the worid

Billabong — Australian term for a riparian wetland that is periodically flooded by the adjacent stream
or river.

Bog — A peat-accumulating wetland that has no significant inflows or outflows and supports acidophilic
mosses, particularly Sphagnum.

Bottomiand — Lowland along streams and rivers, usually on alluvial floodplains, that is periodically
flooded. When forested, it is called a bottomland hardwood forest in the southeastern and eastern
United States.

Carr — Term used in Europe for forested wetlands characterized by alders {Alnus) and willows (Salix).

Cumbungi swamp — Cattail (Typha) marsh in Australia.

Dambo — A seasonally waterlogged and grass-covered linear depression in headwater zone of rivers
with no marked stream channel or woodland vegetation. The term is from the ChiChewa (Central
Africa) dialect meaning “meadow grazing."

Delta — A wetland-river-upland complex located where a river forms distributaries as it merges with the
sea; there are also examples of inland deltas, such as the Peace-Athabasca Delta in Canada and the
Okavango Delta in Botswana (see Chapter 3: “Wetlands of the World”).

Fen — A peat-accumulating wetland that receives some drainage from surrounding mineral soil and
usually supports marshlike vegetation.

Lagoon — Term frequently used in Europe to denote a deepwater enclosed or partially opened aquatic
system, especially in coastal delta regions.

Mangal — Same as mangrove.

Mangrove — Subtropical and tropical coastal ecosystem dominated by halophytic trees, shrubs, and
other plants growing in brackish to saline tidal waters. The word mangrove also refers to the dozens
of tree and shrub species that dominate mangrove wetlands.

Marsh — A frequently or continually inundated wetland characterized by emergent herbaceous
vegetation adapted to saturated soil conditions. In European terminology, a marsh has a mineral soil
substrate and does not accumulate peat. See also tidal freshwater marsh and salt marsh.

Mire — Synonymous with any peat-accumulating wetland (European definition); from the Norse word
myrr. The Danish and Swedish word for peatland is now mose.

Moor — Synonymous with peatland (European definition). A highmoor is a raised bog; a lowmoor is a
peatland in a basin or depression that is not elevated above its perimeter. The primitive sense of the
Old Norse root is "dead” or barren land.

Muskeg — Large expanse of peatlands or bogs; particularly used in Canada and Alaska.

Oxbow — Abandoned river channel, often developing into a swamp or marsh.

Pakihi — Peatland in southwestern New Zealand dominated by sedges, rushes, ferns, and scattered
shrubs. Most pakihi form on terraces or plains of glacial or fluvial outwash origin and are acid and
exceedingly infertile.

Peatland — A generic term of any wetland that accumulates partially decayed plant matter (peat).

Playa — An arid- to semiarid-region wetland that has distinct wet and dry seasons. Term is used for
shallow depressional recharge wetlands occurring in the Great Plains region of North America “that
are formed through a combination of wind, wave, and dissolution processes” (Smith, 2003).

Pocosin — Peat-accumulating, nonriparian freshwater wetland, generally dominated by evergreen
shrubs and trees and found on the southeastern coastal plain of the United States. The term comes
from the Algonquin for “swamp on a hill.”

Pokelogan — Northeastern U.S. marshy or stagnant water that has branched off from a stream or lake.

Pothole — Shallow marshlike pond, particularly as found in the Dakotas and central Canadian
provinces, the so-called prairie pothole region.

Raupo swamp — Cattail (Typha) marsh in New Zealand.

Reedmace swamp — Cattail (Typha) marsh in the United Kingdom.

Reedswamp — Marsh dominated by Phragmites (common reed); term used particularly in Europe.

Wetland Common Terms 35

Table 2.1 (Continued)

Riparian ecosystem — Ecosystem with a high water table because of proximity to an aquatic
ecosystem, usually a stream or river. Also called bottomland hardwood forest, floodplain forest,
bosque, riparian buffer, and streamside vegetation strip.

Salt marsh — A halophytic grassland on alluvial sediments bordering saline water bodies where water
level fluctuates either tidally or nontidally.

Sedge meadow — Very shallow wetland dominated by several species of sedges (e.g., Carex, Scirpus,
Cyperus).

Shrub-Scrub Swamp — A freshwater wetland transitional between a forested swamp and a wet

meadow or marsh, dominated by shrubs, with trees having less than 20 percent cover and less that
10 m height.

Slough — An elongated swamp or shallow lake system, often adjacent to a river or stream. A slowly
flowing shallow swamp or marsh in the southeastern United States (e.g., cypress slough). From the
Old English word s/oh, meaning a watercourse running in a hollow.

Strand — Similar to a slough; a slow-flowing riverine/wetland system, often forested, found especially
in south Florida, where gradients are low.

Swamp — Wetland dominated by trees or shrubs (U.S. definition). In Europe, forested fens and
wetlands dominated by reed grass {Phragmites) are also called swamps (see reedswamp).

Tidal freshwater marsh — Marsh along rivers and estuaries close enough to the coastline to
experience significant tides by nonsaline water. Vegetation is often similar to nontidal freshwater
marshes.

Tbrieugh — Areas seasonally flooded by karst groundwater with sufficient frequency and duration to
produce wetland characteristics. They generally flood in winter and are dry in summer and fill and
empty through underground passages. Term is specific for these types of wetlands found mostly in
western Ireland.

Varzea — A seasonally flooded forest in the Amazon River Basin. It usually refers to forests flooded by
Whitewater (sediment-laden) river water.

Vernal pool — Shallow, intermittently flooded wet meadow, generally typical of Mediterranean climate
with dry season for most of the summer and fall. Term is now used to indicate wetlands temporarily
flooded in the spring throughout the United States.

Viols — Seasonal wetland similar to a dambo; term used in southern Africa.

Wad (pi. waddon) — Unvegetated tidal flat originally referring to the northern Netherlands and
northwestern German coastline. Now used throughout the world for coastal areas.

Wat moadow — Grassland with waterlogged soil near the surface but without standing water for most
of the year.

Wat prairio — Similar to a marsh, but with water levels usually intermediate between a marsh and a
wet meadow.

marshes. European terminology, which is much older, is also much richer and distin-
guishes at least four different kinds of freshwater wetlands — from mineral-rich reed
beds, called reedswamps, to wet grassland marshes, to fens, and, finally, to bogs or
moors. To some, all of these wetland types are considered mires. According to others,
mires are limited to peat-building wetlands. The European classification is based on
the amount of surface water and nutrient inflow (rheotrophy), type of vegetation, pEI,
and peat- building characteristics.

Two points can be made about the use of common terms in classifying wetland
types: First, the physical and biotic characteristics grade continuously from one of these

36 Chapter 2 Wetland Definitions

wetland types to the next; hence, any classification based on common terms is, to an
extent, arbitrary. Second, the same term may refer to different systems in different
regions. The common terms continue to be used, even in the scientific literature;
we simply suggest that they be used with caution and with an appreciation for an
international audience.

Formal Wetland Definitions

Precise wetland definitions are needed for two distinct interest groups: (1) wetland
scientists and (2) wedand managers and regulators. The wetland scientist is inter-
ested in a flexible yet rigorous definition that facilitates classification, inventory, and
research. The wetland manager is concerned with laws or regulations designed to pre-
vent or control wetland modification and, thus, needs clear, legally binding definitions.
Because of these differing needs, different definitions have evolved for the two groups.
The discrepancy between the regulatory definition o? jurisdictional wetlands other
definitions in the United States has meant, for example, that maps developed for wet-
land inventory purposes cannot be used for regulating wetland development. This is
a source of considerable confusion to regulators and landowners.

Definitions that are more scientific in nature are presented in this section. Defini-
tions that are more used in a legal sense are presented in the next section.

Early U.S. Definition: Circular 39 Definition

One of the earliest definitions of the term wetlands presented by the U.S. Fish
and Wildlife Service in 1956 in a publication that is frequendy referred to as Circular
39 (Shaw and Fredine, 1956):

The term “wedands” . . . refers to lowlands covered with shallow and sometimes
temporary or intermittent waters. They are referred to by such names as marshes,
swamps, bogs, wet meadows, potholes, sloughs, and river-overflow lands. Shallow
lakes and ponds, usually with emergent vegetation as a conspicuous feature, are
included in the definition, but the permanent waters of streams, reservoirs, and deep
lakes are not included. Neither are water areas that are so temporary as to have little
or no effect on the development of moist-soil vegetation.

The Circular 39 definition ( 1 ) emphasized wetlands that were important as water-
fowl habitats and (2) included 20 types of wetlands that served as the basis for the main
wetland classification used in the United States until the 1970s (see Chapter 13). It
thus served the limited needs of both wetland managers and wetland scientists.

U.S. Fish and Wildlife Service Definition

Perhaps the most comprehensive definition of wetlands was adopted by wetland sci-
entists in the U.S. Fish and Wildlife Service in 1979, after several years of review. The

Formal Wetland Definitions 37

definition was presented in a report entided Classification of Wetlands and Deepwater
Habitats of the United States (Cowardin et al., 1979):

Wetlands are lands transitional between terrestrial and aquatic systems where the
water table is usually at or near the surface or the land is covered by shallow
water .... Wetlands must have one or more of the following three attributes: (1) at
least periodically, the land supports predominantly hydrophytes; (2) the substrate is
predominantly undrained hydric soil; and (3) the substrate is nonsoil and is saturated
with water or covered by shallow water at some time during the growing season of
each year.

This definition was significant for its introduction of several important concepts
in wetland ecology. It was one of the first definitions to introduce the concepts of
hydric soils and hydrophytes, and it served as the impetus for scientists and managers to
define these terms more accurately (National Research Council, 1995). Designed tor
scientists as well as managers, it is broad, flexible, and comprehensive, and includes
descriptions of vegetation, hydrology, and soil. It has its main utility in scientific studies
and inventories and generally has been more difficult to apply to the management and
regulation of wetlands. It is still frequently accepted and employed in the United States
today and was, at one time, accepted as the official definition of wetlands by India.
Like the Circular 39 definition, this definition serves as the basis for a detailed wetland
classification and an updated and comprehensive inventory of wetlands in the United
States. The classification and inventory are described in more detail in Chapter 13.

Canadian Wetland Definitions

Canadians, who deal with vast areas of inland northern peatlands, have developed
a specific national definition of wetlands. Two definitions were formally published in
the book Wetlands of Canadahy the National Wetlands Working Group (1988). First,
Zoltai (1988) defined a wetland as:

Land that has the water table at, near, or above the land surface or which is saturated
for a long enough period to promote wetland or aquatic processes as indicated by
hydric soils, hydrophytic vegetation, and various kinds of biological activity which
are adapted to the wet environment.

Zoltai (1988) also noted that “wetlands include waterlogged soils where in
some cases the production of plant materials exceeds the rate of decomposition.” He
describes the wet and dry extremes of wetlands as:

■ Shallow open waters, generally less than 2 m; and

■ Periodically inundated areas only if waterlogged conditions dominate
throughout the development of the ecosystem.

Tarnocai et al. (1988) offered a slightly reworded definition in that same pub-
lication as the basis of the Canadian wetland classification system. That definition,

38 Chapter 2 Wetland Definitions

repeated by Zoltai and Vitt (1995) and Warner and Rubec (1997) in later years,
remains the official definition of wedands in Canada:

Land that is saturated with water long enough to promote wetland or aquatic
processes as indicated by poorly drained soils, hydrophytic vegetation and various
kinds of biological activity which are adapted to a wet environment.

These definitions emphasize wet soils, hydrophytic vegetation, and “various
kinds” of other biological activity. The distinction between “hydric soils” in the
Zoltai definition and “poorly drained soils” in the current, more accepted definition
may be a reflection of the reluctance by some to use hydric soils exclusively to define
wetlands. Hydric soils are discussed in more detail in Chapter 5: “Wetland Soils.”

U.S. National Academy of Sciences Definition

In the early 1990s, amid renewed regulatory controversy in the United States as to
what constitutes a wetland, the U.S. Congress asked the private, nonprofit National
Academy of Sciences to appoint a committee through its principal operating agency,
the National Research Council (NRC), to undertake a review of the scientific aspects of
wetland characterization. The committee was charged with considering: (1) the ade-
quacy of the existing definition of wetlands; (2) the adequacy of science for evaluating
the hydrologic, biological, and other ways that wetlands function; and (3) regional
variation in wedand definitions. The report produced by that committee two years
later was entitled Wetlands: Characteristics and Boundaries 1995) and included

yet another scientific definition, referred to as a “reference definition” in that it was
meant to stand “outside the context of any particular agency, policy or regulation”:

A wetland is an ecosystem that depends on constant or recurrent, shallow inundation
or saturation at or near the surface of the substrate. The minimum essential
characteristics of a wetland are recurrent, sustained inundation or saturation at or
near the surface and the presence of physical, chemical, and biological features
reflective of recurrent, sustained inundation or saturation. Common diagnostic
features of wetlands are hydric soils and hydrophytic vegetation. These features will
be present except where speciflc physiochemical, biotic, or anthropogenic factors
have removed them or prevented their development.

Although little formal use has been made of this definition, it remains the most
comprehensively developed scientific wedand definition. It uses the terms hydric soils
and hydrophytic vepfetation^ as did the early U.S. Fish and Wildlife Service definition,
but indicates that they are “common diagnostic features” rather than absolute neces-
sities in designating a wetland.

An International Definition

The International Union for the Conservation of Nature and Natural Resources
(lUCN) at the Convention on Wetlands of International Importance Especially as
Waterfowl Habitat, better known as the Ramsar Convention, adopted the following

Legal Definitions 39

definition of wetlands in Article 1.1 of the Convention of Wedands (Finlayson and
Moser, 1991):

For the purposes of this Convention wetlands are areas of marsh, fen, peatland or
water, whether natural or artificial, permanent or temporary, with water that is static
or flowing, fresh, brackish, or salt including areas of marine water, the depth of
which at low tide does not exceed six meters.

The Ramsar definition further provides in Article 1.2 of the Convention that
wetlands

may incorporate riparian and coastal zones adjacent to the wetlands, and islands or
bodies of marine water deeper than six metres at low tide lying within the wetlands.

This definition, which was adopted at the first meeting of the convention in Ram-
sar, Iran, in 1971, does not include vegetation or soil and extends wetlands to water
depths of 6 meters or more, well beyond the depth usually considered wetlands in
the United States and Canada. The rationale for such a broad definition of wetlands
“stemmed from a desire to embrace all the wetland habitats of migratory water birds”
(Scott and Jones, 1995).

Legal Definitions

When protection of wetlands began in earnest in the mid-1970s in the United States,
there arose an almost immediate need for precise definitions that were based as much
on closing legal loopholes as on science. Two such definitions have developed in U.S.
agencies — one for the U.S. Army Corps of Engineers to enforce its legal responsibil-
ities with a “dredge-and-fill” permit program in the Clean Water Act and the other
for the U.S. Natural Resources Conservation Service to administer wetland protection
under the so-called swampbuster provision of the Food Security Act. Both agencies
were parties to an agreement in 1993 to work together on administering a unified
policy of wetland protection in the United States, yet the two separate definitions
remain.

U.S. Army Corps of Engineers Definition

A U.S. government regulatory definition of wetlands is found in the regulations used
by the U.S. Army Corps of Engineers for the implementation of a dredge-and-fill
permit system required by Section 404 of the 1977 Clean Water Act amendments.
That definition has now survived several decades in the legal world and is given as
follows:

The term “wetlands” means those areas that are inundated or saturated by surface or
ground water at a frequency and duration sufficient to support, and that under
normal circumstances do support, a prevalence of vegetation typically adapted for life
in saturated soil conditions. Wetlands generally include swamps, marshes, bogs, and
similar areas. (33 CFR 328.3(b); 1984)

40 Chapter 2 Wetland Definitions

This definition replaced a 1975 definition that stated “those areas that normally
are characterized by the prevalence of vegetation that saturated soil conditions

tor growth and reproduction” (42 Fed. Rejj. 3712X, July 19, 1977; italics added),
because the Corps of Engineers found that the old definition excluded “many forms
of truly aquatic vegetation that are prevalent in an inundated or saturated area, but
that do not require saturated soil from a biological standpoint for their growth and
reproduction.” The words normally in the old definition and that under normal cir-
cumstances do support in the new definition were intended “to respond to situations
in which an individual would attempt to eliminate the permit review requirements of
Section 404 by destroying the aquatic vegetation” (quotes from 42 Fed. Rejj. 37128,
July 19, 1977). The need to revise the 1975 definition illustrates how difficult it has
been to develop a legally useful definition that also accurately reflects the ecological
reality of a wetland site.

This legal definition of wetlands has been debated in the courts in several cases,
some of which have become landmark cases. In one of the first court tests of wetland
protection, the Fifth Circuit of the U.S. Court of Appeals ruled in 1972, in Zabel v.
Tabb, that the U.S. Army Corps of Engineers has the right to refuse a permit for filling
of a mangrove wetland in Florida. In 1975, in Natural Resources Defense Council v.
Callaway., wetlands were included in the category “waters of the United States,” as
described by the Clean Water Act. Prior to that time, the Corps of Engineers regulated
dredge-and-fill activities (Section 404 of the Clean Water Act) for navigable waterways
only; since that decision, wetlands have been legally included in the definition of waters
of the United States.

In 1985, the question of regulation of wetlands reached the U.S. Supreme Court
for the first time. The court upheld the broad definition of wetlands to include
groundwater-fed wetlands in United States v. Riverside Bayview Homes, Inc. In that
case, the Supreme Court affirmed that the U.S. Army Corps of Engineers had
jurisdiction over wetlands that were adjacent to navigable waters, but it left open the
question as to whether it had jurisdiction over nonadjacent wetlands (NRC, 1995).
The legal definition of wetlands has been involved in the U.S. Supreme Court three
times more, m 2001, 2006, and 2013; these cases are discussed in more detail in
Chapter 15 “Wetland Laws and Protection.”

Food Security Act Definition

In December 1985, the U.S. Department of Agriculture, through its Soil Conserva-
tion Service [now known as the Natural Resources Conservation Service (NRCS)],
was brought into the arena of wetland definitions and wetland protection by means
of a provision known as swampbuster in the 1985 Food Security Act. On agricultural
land in the United States that, prior to December 1985, had been exempt from reg-
ulation, wetlands were now protected. As a result of this swampbuster provision, a
definition, known as the NRCS or Food Security Act definition, was included in the
Act (16 CFR801(a)(16); 1985):

Legal Definitions 41

The term “wetland” except when such term is part of the term “converted wetland”
means land that —

(A) has a predominance of hydric soils;

(B) is inundated or saturated by surface or ground water at a frequency and
duration sufficient to support a prevalence of hydrophytic vegetation typically
adapted for life in saturated soil conditions; and

For purposes of this Act and any other Act, this term shall not include lands in Alaska
identified as having high potential for agricultural development which have a
predominance of permafrost soils.

The emphasis on this agriculture-based definition is on hydric soils. The omission
of wetlands that do not have hydric soils, while not invalidating this definition, makes
it less comprehensive than some others — for example, the NRC (1995) definition.
A curious feature of this definition is its wholesale exclusion of the largest state in
the United States from the definition of wetlands. The exclusion of Alaskan wetlands
that have a high potential for agriculture makes this definition even less of a scientific
and more of a regulatory or even political definition. There is no scientific distinction
between the characteristics of Alaskan wedands and wetlands in the rest of the United
States except for climatic differences and the presence of permafrost under many but
certainly not all Alaskan wetlands (NRC, 1995).

Jurisdictional Wetlands

Since 1989, the term jurisdictional wetland has been used for legally defined wedands
in the United States to delineate those areas that are under the jurisdiction of Section
404 of the Clean Water Act or the swampbuster provision of the Food Security Act.
The U.S. Army Corps of Engineers’ definition cited previously emphasizes only one
indicator, vegetadve cover, to determine the presence or absence of a wedand. It is
difficult to include soil informadon and water condidons in a wedand definidon when
its main purpose is to determine jurisdicdon for regulatory purposes and there is litde
dme to examine the site in detail. The Food Security Act definition, however, includes
hydric soils as the principal determinant of wetlands.

It is likely that most of the wedands that are considered jurisdictional wedands by
the preceding two legal definidons fit the scientific definition of wetlands. It is also just
as likely that some types of wetlands, particularly those that have less chance of devel-
oping hydric soil characteristics or hydrophytic vegetadon (e.g., riparian wedands),
would not be idendfied as jurisdicdonal wedands with the legal definidons. And of
course, excluding Alaskan wedands “having high potential for agricultural develop-
ment” from the Food Security Act definidon has no scientific basis at all but is a
polidcal decision.

42 Chapter 2 Wetland Definitions

Those who delineate wedands are interested in a definition that allows the rapid
identification of a wetland and the degree to which it has been or could be altered.
They are interested in the delineation of wetland boundaries, and establishing bound-
aries is facilitated by defining the wetland simply, according to the presence or absence
of certain species of vegetation or aquatic life or the presence of simple indicators,
such as hydric soils. Several U.S. federal manuals spelling out specific methodolo-
gies for identifying jurisdictional wetlands were written or proposed in the 1980s and
early 1990s. The manuals differed, however, in the prescribed ways these three crite-
ria are proved in the field. The first of these manuals (U.S. Army Corps of Engineers,
1987), is now the accepted version and is widely used to field -identify wetlands in the
United States. All three manuals indicated that the three criteria for wetlands — namely,
wetland hydrology, wetland soils, and hydrophytic vegetation — must be present. As
illustrated in Figure 2.2, these three variables are not independent; strong evidence of
long-term wetland hydrology, for example, should almost ensure that the other two
variables are present. Furthermore, potentially other indicators of the physiochemistry
and biota beyond hydric soils and hydrophytic vegetation may one day serve as useful
indicators of wetlands.

Choice of a Definition

A wetland definition that will prove satisfactory to all users has not yet been devel-
oped because the definition of wetlands depends on the objectives and the field of
interest of the user. Different definitions can be formulated by the geologist, soil
scientist, hydrologist, biologist, ecologist, sociologist, economist, political scientist,
public health scientist, and lawyer. This variance is a natural result of the differences
in emphasis in the definer’s training and of the different ways in which individual dis-
ciplines deal with wetlands. For ecological studies and inventories, the U.S. Fish and
Wildlife Service definition has been and should continue to be applied to wetlands
in the United States. Although somewhat generous in defining wetlands on the wet
edge, the Ramsar definition is firmly entrenched in international circles. When wetland
management, particularly regulation, is necessary, the U.S. Army Corps of Engineers’
definition, as modified, is probably most appropriate.

Just as important as the precision of the definition of a wetland, however, is the
consistency with which it is used. That is the difficulty we face when science and
legal issues meet, as they often do, in resource management questions such as wet-
land conservation versus wetland drainage. Applying a comprehensive definition in a
uniform and fair way requires a generation of well-trained wetland scientists and man-
agers armed with a fundamental understanding of the processes that are important
and unique to wetlands.

Recommended Readings

Fraser, L. H., and P. A. Keddy, eds. 2005. The World’s Largest Wetlands: Ecology and
Conservation. Gambridge, UK: Gambridge University Press.

References 101

Grunwald, M. 2006. The Swamp: The Everglades, Florida, and the Politics of Paradise.
New York: Simon & Schuster.

Junk, W., ed. 2013. The World’s Wetlands and Their Future under Global Glimate
Ghange. Special issue of Aquatic Sciences 75 (1): 1-167.

References

Abraham, K. F. and G. J. Keddy. 2005. The Hudson Bay Lowland. In L.A. Fraser
and P.A. Keddy, eds.. The World’s Larpiest Wetlands: Ecology and Conservation,
Gambridge University Press, Gambridge, UK, pp. 118-148.

Altinbilek, D. 2004. Development and management of the Euphrates-Tigris basin.
Water Resources Development 20: 15-33.

Aselmann, 1., and P. J. Grutzen. 1989. Global distribution of natural freshwater wet-
lands and rice paddies, their net primary productivity, seasonality and possible
methane emissions. Journal of Atmospheric Chemistry 8: 307-358.

Australian Nature Gonservation Agency. 1996. Wetlands Are Important. Two-page
flyer. National Wetlands Program, ANGA, Ganberra, Australia.

Bartlett, G. H. 1904. Tales of Kankakee Land. Gharles Scribner’s Sons, New York.
232 pp.

Bazilevich, N. 1., L. Ye. Rodin, and N. N. Rozov. 1971. Geophysical aspects of bio-
logical productivity. Soviet Geography 12: 293-317.

Bruins, R. J. F., S. Gai, S. Ghen, and W. J. Mitsch. 1998. Ecological engineering
strategies to reduce flooding damage to wetland crops in central Ghina. Ecolopfical
Engineering 11: 231-259.

Galheiros, D. F., M. D. de Oliveira, and G. R. Padovani. 2012. Hydro-ecological pro-
cesses and anthropogenic impacts on the ecosystem services of the Pantanal wet-
land. In A. A. R. loris, ed.. Tropical Wetland Management: The South-American
Pantanal and the International Experience . Ashgate Publishing Ltd., Farnham,
England, pp. 29-57.

Gampbell, D. 2005. The Gongo River basin. In L.A. Fraser and P.A. Keddy, eds.. The
World’s Larpiest Wetlands: Ecology and Conservation, Gambridge University Press,
Gambridge, UK, pp. 149-165.

Ghambers, J. M., and A. J. McGomb. 1994. Establishment of wetiand ecosystems in
lakes created by mining in Western Australia. In W. J. Mitsch, ed. , Global Wetlands:
Old World and New. Elsevier, Amsterdam, The Netherlands, pp. 431-441.

Ghen, Y. 1995. Study of Wetlands in China. Jilin Sciences Technology Press,
Ghangchun, Ghina. [in Ghinese with English summaries.]

Glarkson, B. R. 1997. Vegetation recovery following fire in two Waikato peatlands at
Whangamarino and Moanatuatua. New Zealand Journal of Botany 35: 167-179.

Gomin, F. A., J. A. Romero, V. Astorga, and G. Garcia. 1997. Nitrogen removal and
cycling in restored wetlands used as filters of nutrients for agricultural runoff.
Water Science and Technolopiy 35: 255-261.

102 Chapter 3 Wetlands of the World

Contreras-Espinosa, F., and B. G. Warner. 2004. Ecosystem characteristics and
management considerations from coastal wetlands in Mexico. Hydrobiologia 511:
233-245.

Costanza, R., W. J. Mitsch, and J. W. Day. 2006. A new vision for New Orleans and
the Mississippi delta: Applying ecological economics and ecological engineering.
Frontiers in Ecology and the Environment 4: 465-472.

Dahl, T. E. 1990. Wetlands losses in the United States, 1780s to 1980s. U.S. Depart-
ment of Interior, Fish and Wildlife Service, Washington, DC. 21 pp.

Dahl, T. E. 2000. Status and trends of wetlands in the conterminous United States
1986 to 1997. U.S. Department of the Interior, Fish and Wildlife Service, Wash-
ington, DC, 82 pp.

Dahl, T. E. 2006. Status and trends of wetlands in the conterminous United States
1998 to 2004. U.S. Department of the Interior, Fish and Wildlife Service, Wash-
ington, DC, 112 pp.

Dahl, T. E. 2011. Status and trends of wetlands in the conterminous United States
2004 to 2009. U.S. Department of the Interior, Fish and Wildlife Service, Wash-
ington, DC, 108 pp.

Dahl, T. E. 2014. Status and trends of prairie wetlands in the United States 1997 to
2009. U.S. Department of the Interior, Fish and Wildlife Service, Washington,
DC, 67 pp.

Dahl, T. E., and C. E. Johnson. 1991 . Wetlands status and trends in the conterminous
United States mid-1970s to mid-1980s. U.S. Department of Interior, Fish and
Wildlife Service, Washington, DC. 28 pp.

da Silva, C. J., and P. Girard. 2004. New challenges in the management of the Brazilian
Pantanal and catchment area. Wetlands Ecology and Management 12: 553-561.

Davidson, N.C. 2014 How much wetland has the world lost.^ Long-term and recent
trends in global wetland area. Marine and Freshwater Research 65: 934-941.

Day, J. W., Jr., J. Barras, E. Clairain, J. Johnston, D. Justic, G. P. Kemp, J.-Y. Ko, R.
Lane, W. J. Mitsch, G. Steyer, P. Templet, and A. Yahez-Arancibia. 2005. Implica-
tions of global climatic change and energy cost and availability for the restoration
of the Mississippi Delta. Ecological Engineering 24: 253-265.

Day, J. W, Jr., D. F. Boesch, E. J. Clairain, G. P. Kemp, S. B. Laska, W. J. Mitsch, K.
Orth, H. Mashriqui, D. R. Reed, L. Shabman, C. A. Simenstad, B. J. Streever,
R. R. Twilley, C. C. Watson, J. T. Wells, and D. F. Whigham. 2007. Restoration
of the Mississippi Delta: Lessons From Hurricanes Katrina and Rita. Science 315:
1679-1684.

Dennis, J. V. 1988. The Great Cypress Swamps. Louisiana State University Press, Baton
Rouge. 142 pp.

Douglas, M. S. 1947. The Everpflades: River of Grass. Ballantine, New York. 308 pp.

Dugan, P. 1993. Wetlands in Danger. Michael Beasley, Reed International Books,
London. 192 pp.

Ellison, A. M. 2004. Wetlands of Central America. Wetlands Ecology and Management
12: 3-55.

References 103

Finlayson, M., and N. C. Davidson. 1999. Global Review of Wetiand Resources and
Priorities for Wetland Inventory. Ramsar Bureau Contract 56, Ramsar Convention
Bureau, Gland, Switzerland.

Forsyth, J. L. 1960. The Black Swamp. Ohio Department of Natural Resources, Divi-
sion of Geological Survey, Columbus, OH. 1 p.

Frayer, W. E., T. J. Monahan, D. C. Bowden, and F. A. Graybill. 1983. Status and
trends of wetlands and deepwater habitat in the conterminous United States,
1950s to 1970s. Department of Forest and Wood Sciences, Colorado State Uni-
versity, Fort Collins. 32 pp.

Gorham, E. 1991. Northern peatlands: Role in the carbon cycle and probable
responses to climatic warming. Ecological Applications 1: 182-195.

Grant, M. 1962. Myths of the Greeks and Romans. Mentor/New American Library,
New York.

Gray, L. C., O. E. Baker, F. J. Marschner, B. O. Weitz, W. R. Chapline, W. Shepard,
and R. Zon. 1924. The utilization of our lands for crops, pasture, and forests.
In U.S. Department of Agriculture Yearbook., 1923. Government Printing Office,
Washington. DC, pp. 415-506.

Grunwald, M. 2006. The Swamp: The Everpflades, Florida, and the Politics of Paradise.
Simon & Schuster, New York, 450 pp.

Hanby, J., and D. Bygott. 1998. N^oron^oro Conservation Area. Kibuyu Partners,
Karatu, Tanzania.

Harper, D. M., C. Adams, and K. Mavuti. 1995. The aquatic plant communities of
the Lake Naivasha wetland, Kenya: Pattern, dynamics, and conservation. Wetlands
Ecology and Management 3: 111-123.

Harris, M. B., W. Tomas, G. Mourao, C. J. DaSilva, E. Guimaraes, F. Sonoda, and
E. Fachim. 2005. Safeguarding the Pantanal wetlands: Threats and conservation
initiatives. Conservation Biology 19: 714-720.

loris, A. A. R., ed. 2012. Tropical Wetland Management: The South-American Pan-
tanal and the International Experience. Ashgate Publishing Ltd., Farnham, Eng-
land, 351 pp.

Johnson, D. C. 1942. The Origin of the Carolina Bays. Columbia University Press,
New York. 341 pp.

Johnson, P, and P. Gerbeaux. 2004. Wetland Types in New Zealand. New Zealand
Department of Conservation, Wellington, 184 pp.

Junk, W. J. 1993. Wetlands of tropical South America. In D. F. Whigham,
D. Dykyjova, and S. Hejny, eds.. Wetlands of the World, I: Inventory, Ecolojjy,
and Manapfement. Kluwer Academic Publishers, Dordrecht, The Netherlands,
pp. 679-739.

Junk, W. J., and M. T. F. Piedade. 2004. Status of knowledge, ongoing research, and
research needs in Amazonian wetlands. Wetlands Ecology and Management 12:
597-609.

Junk, W. J., and M. T. F. Piedade. 2005. The Amazon River basin. In L.A. Fraser
and PA. Keddy, eds.. The WorldS Larjjest Wetlands: Ecology and Conservation,
Cambridge University Press, Cambridge, UK, pp. 63-117.

104 Chapter 3 Wetlands of the World

Junk, W. J., and C. Nunes de Cunha. 2005. Pantanal: A large South American wetland
at a crossroads. Ecological En£iineerin£i 24: 391-401.

Kaatz, M. R. 1955. The Black Swamp: A study in historical geography. Annals of the
Association of American Geographers 35: 1-35.

Kirk, R W., Jr. 1979. The Great Dismal Swamp. University Press of Virginia, Char-
lottesville. 427 pp.

Lehner, B. and P Doll. 2004. Development and validation of a global database of
lakes, reservoirs, and wetlands. Journal of Hydrology 296: 1-22.

Tide, R. F., V. G. Meentemeyer, J. E. Pinder, and L. M. Beatty 1995. Hydrology of
a Carolina Bay located on the Upper Coastal Plain of western South Carolina.
Wetlands 15: 47-57.

Lodge, T. E. 2010. The Everglades Handbook: Understanding the Ecosystem., 3rd ed.
CRC Press, Boca Raton, EL.

Lu, J., ed. 1990. Wetlands in China. East China Normal University Press, Shanghai.
177 pp. [in Chinese].

Lu, J. 1995. Ecological significance and classification of Chinese wetlands. Vegetatio
118: 49-56.

Ma, X., X. Liu, and R. Wang. 1993. China’s wetlands and agro-ecological engineering.
Ecological Engineering 2: 291-330.

Maltby, E., and R. E. Turner. 1983. Wetlands of the world. Geographic Magazine 55:
12-17.

Matthews, E., and 1. Fung. 1987. Methane emissions from natural wetlands: Global
distribution, area, and environmental characteristics of sources. Global Biogeo-
chemical Cycles 1: 61-86.

Mauchamp, A., P. Chauvelon, and P. Grillas. 2002. Restoration of floodplain wedands:
Opening polders along a coastal river in Mediterranean France, Vistre marshes.
Ecological Engineering 18: 619-632.

McCarthy, J. M., T. Gumbricht, T. McCarthy, P Frost, K. Wessels, and F. Seidel.
2004. Flooding patterns in the Okavango wetland in Botswana between 1972
and 2000. Ambio 32: 453-457.

Mendelsohn, J. and S. el Obeid. 2004. Okavango River: The Flow of a Lifeline. Struik
Publishers, Cape Town, South Africa. 176 pp.

Meyer, A. H. 1935. The Kankakee “Marsh” ofnorthern Indiana and Illinois. Michigan
Academy of Science, Arts, and Letters Papers 21: 359-396.

Michalak, A. M., E. J. Anderson, D. Beletsky, S. Boland et al. 2013. Record-setting
algal bloom in Lake Erie caused by agricultural and meteorological trends con-
sistent with expected future conditions. Proceedings National Academy of Sciences
110: 6524-6529.

Millennium Ecosystem Assessment, 2005. Ecosystems and Human Well-being: Syn-
thesis. Island Press and World Resources Institute, Washington, DC. 137 pp.

Minnemeyer, S. 2002. An Analysis of Access into Central Africa’s Rainforests . World
Resources Institute, Washington DC.

References 105

Mitsch, W. J. 2010. The 2010 Gulf of Mexico oil spill: What would Mother Nature
do.> Ecological En£iineerin£i 36: 1607-1610.

Mitsch, W. J., R. H. Mitsch, and R. E. Turner. 1994. Wetlands of the Old and New
Worlds — Ecology and Management. In W. J. Mitsch, ed.. Global Wetlands: Old
and New. Elsevier, Amsterdam, The Netherlands, pp. 3-56.

Mitsch, W. J., A. M. Nahlik, R Wolsld, B. Bernal, L. Zhang, and L. Ramberg.
2010. Tropical wedands: Seasonal hydrologic pulsing, carbon sequestration, and
methane emissions. Wetlands Ecology and Management 18: 573-586.

Mitsch, W. J, and M. Hernandez. 2013. Landscape and climate change threats to
wetlands of North and Central America. Aquatic Sciences 75: 133-149.

Nahlik, A. M., and W. J. Mitsch. 2006. Tropical treatment wetlands dominated by
free-floating macrophytes for water quality improvement in the Caribbean coastal
plain of Costa Rica. Ecological Engineering 28: 246-257.

National Wetlands Working Group. 1988. Wetlands of Canada. Ecological and
Classification Series 24, Environment Canada, Ottawa, Ontario, and Polyscience
Publications, Montreal, Quebec. 452 pp.

Niu, Z., H. Zhang, and P. Gong. 2011. More protection for China’s wetlands. Nature
471: 305.

Novacek, J. M. 1989. The water and wetland resources of the Nebraska Sandhills. In
A. G. van der Valk, ed.. Northern Prairie Wetlands. Iowa State University Press,
Ames, pp. 340-384.

Odum, H. T. 1951. The Carolina Bays and a Pleistocene weather map. American
Journal of Science 250: 262-270.

Perez-Arteaga, A., K. J. Gaston, and M. Kershaw. 2002. Undesignated sites in Mexico
qualifying as wedands of international importance. Biological Conservation 107:
47-57.

Por, F. D. 1995. The Pantanal of Mato Grosso (Brazil). Kluwer Academic Publishers,
Dordrecht, The Netherlands. 122 pp.

Prasad, V. P, D. Mason, J. E. Marburger, and C. R. A. Kumar. 1996. Illustrated
Flora ofKeoladeo National Park, Bharatpur, Rajasthan. Bombay Natural History
Society, Mumbai, India. 435 pp.

Prigent, C., F. Papa, F. Aires, C. Jimenez, W. B. Rossow, and E. Matthews. 2012.
Changes in land surface water dynamics since the 1990s and relation to population
pressure. Geophys. Res. Lett., doi:10.1029/2012GL051276

Prouty, W. F. 1952. Carolina Bays and their origin. Geological Society of America Bul-
letin 63: 167-224.

Qiu, J. 2011. China faces up to ‘“terrible”’ state of its ecosystems. Nature 471: 19.

Ramberg, L., P. Wolsld, and M. Krah. 2006a. Water balance and infiltration in a sea-
sonal floodplain in the Okavango Delta, Botswana. Wetlands 26: 677-690.

Ramberg, L., P. Hancock, M. Lindholm, T. Meyer, S. Ringrose, J. Silva., J. Van As,
and C. VanderPost. 2006b. Species diversity of the Okavango Delta, Botswana.
Aquatic Sciences 6&\ 310-337.

106 Chapter 3 Wetlands of the World

Ramsar Convention Secretariat. 2004. Ramsar Handbook for the Wise Use of Wetlands^
2nd ed. Handbook 10, Wetland Inventory: A Ramsar framework for wetland
inventory. Ramsar Secretariat, Gland, Switzerland.

Revenga, C., J. Brunner, N. Henninger, K. Kassem, and R. Payne. 2000. Pilot Analysis
of Global Ecosystems: Freshwater Systems. World Resources Institute, Washington,
DC. 65 pp.

Richardson, C. J., R. Evans, and D. Carr. 1981. Pocosins: An ecosystem in transi-
tion. In C. J. Richardson, ed., Pocosin Wetlands. Hutchinson Ross Publishing,
Stroudsburg, PA, pp. 3-19.

Ringrose, S., C. Vanderpost, V. Matheson, P. Wolski, P. Huntsman-Mapila,
M. Murray-Hudson, and A. Jellema. 2007. Indicators of desiccation-driven
change in the distal Okavango Delta, Botswana. Journal of Arid Environments
68: 88-112.

Roe, H. B., and Q. C. Ayres. 1954. En^tineerin^ for Agricultural Drainage.
McGraw-Hill, New York. 501 pp.

Russi D., P. ten Brink, A. Farmer, T. Badura, D. Coates, J. Forster, R. Kumar, and N.
Davidson. 2013. The Economics of Ecosystems and Biodiversity for Water and
Wetlands. lEEP, Ramsar Secretariat, Gland, London and Brussels.

Sabin, T.J. and V.T. Holliday. 1995. Playas and lunettes on the Southern High Plains:
Morphometric and spatial relationships. Annals of the Association American Geog-
raphers 85: 286-305.

Savage, H. 1983. The Mysterious Carolina Bays. University of South Carolina Press,
Columbia. 121 pp.

Scavia, D., J. D. Allan, K. K. Arend, S. Bartell et al. 2014. Assessing and addressing
the re-eutrophication of Lake Erie: Central basin hypoxia. Journal of Great Lakes
Research 40: 226-246.

Schmidt, K. F. 2001. A true-blue vision for the Danube. Science 294: 1444-1447.

Shaw, S. P, and C. G. Fredine. 1956. Wetlands of the United States, Their Extent, and
Their Value for Waterfowl and Other Wildlife. Circular 39, U.S. Fish and Wildlife
Service, U.S. Department of Interior, Washington, DC. 67 pp.

Shearer, J. C., and B. R. Clarkson. 1998. Whangamarino wetland: Effects of lowered
river levels on peat and vegetation. International Peat Journal 8: 52-65.

Sheffield, R. M., T. W. Birch, W. H. McWilliams, and J. B. Tansey. 1998. Chamaecy-
paris thyoides (Atlantic white cedar) in the United States. In A. D. Laderman, ed.,
Coastally Restricted Forests. Oxford University Press, New York, pp. 111-123.

Shiel, R. J. 1994. Death and life of the billabong. In X. Collier, ed.. Restoration of
Aquatic Habitats. Selected Papers from New Zealand Limnological Society 1993
Annual Conference, Department of Conservation, pp. 19-37.

Smith, L.M. 2003. Playas of the Great Plains. University of Texas Press, Austin, Texas,
257 pp.

Solomeshch, A. I. 2005. The West Siberian Lowland. In L. A. Fraser and P. A. Keddy,
eds.. The World’s Largest Wetlands: Ecology and Conservation, Cambridge Univer-
sity Press, Cambridge, UK, pp. 11-62.

References 107

Tiner, R. W. 1984. Wetlands of the United States: Current Status and Recent Trends.
National Wetiands Inventory, U.S. Fish and Wildlife Service, Washington, DC.
58 pp.

Trama F. A., F. L. Riso-Patron, A. Kumar, E. Gonzalez, D. Somma, and M.B. McCoy.
2009. Wetland cover types and plant community changes in response to cat-
tail control activities in the Palo Verde Marsh, Costa Rica. Ecological Restoration
27:278-289.

United Nations Environmental Programme (UNEP). 2001. The Mesopotamian
Marshlands: Demise of an Ecosystem (UNEP/DEWA/TR.01-3 Rev.l), UNEP,
Nairobi, Kenya.

Vasander, H., E.-S. Tuittila, E. Lode., L. Lundin, M. Hornets, T. Sallantaus,
R. Heikkila, M.-L. Pitkanen, and J. Laine. 2003. Status and restoration of
peatlands in northern Europe. Wetlands Ecology and Management 11: 51-63.

Verhoeven, J. T. A. 2014. Wetlands in Europe: Perspectives for restoration of a lost
paradise. Ecoloptical Engineering 66: 6-9.

Vidy, G. 2000. Estuarine and mangrove systems and the nursery concept: which is
which. The case of the Sine Saloum system (Senegal). Wetlands Ecology and Man-
agement 8: 37-51.

Whooten, H. H., and M. R. Purcell. 1949. Farm Land Development: Present and
Future by Clearing, Drainapfe, and Irrigation. Circular 825, U.S. Department of
Agriculture, Washington, DC.

Wolff, W. J. 1993. Netherlands — Wetlands. In E. P. H. Best and J. P. Bakker, eds.,
Netherlands — Wetlands. Kluwer Academic Publishers, Dordrecht, The Nether-
lands, pp. 1-14.

Wright, J. O. 1907. Swamp and Overflow Lands in the United States. Circular 76, U.S.
Department of Agriculture, Washington, DC.

Zoltai, S. C., S. Taylor, J. K. Jeglum, G. F. Mills, and J. D. Johnson. 1988. Wetlands of
boreal Canada. In National Wetlands Working Group, ed.. Wetlands of Canada.
Ecological Land Classification Series 24, Environment Canada, Ottawa, Ontario,
and Polyscience Publications, Montreal, Quebec, pp. 97-154.

Part II

The Wetland Environment

Chapter 4

Wetland Hydrology

Hydrologic conditions are extremely important for the maintenance of a
wetland’s structure and function. They affect many abiotic factors, includinp;
soil anaerobiosis, nutrient availability, and, in coastal wetlands, salinity. These,
in turn, determine the biota that develops in a wetland. Finally, completinpi the
cycle, biotic components are active in alterinp; the wetland hydrology and other
physicochemical features. The hydroperiod, or hydrologic signature of a wetland,
is the result of the balance between inflows and outflows of water ( called the
water budget), the wetland basin geomorphology, and subsurface conditions.

The hydroperiod can have dramatic seasonal and year-to-year variations, yet it
remains the major determinant of wetland processes. The major components of
a wetland’s water budget include precipitation, evapotranspiration, surface
inflows and outflows including overbank flooding into riparian wetlands,
groundwater fluxes, and tides or seiches in coastal wetlands. Simple
determinations of the hydroperiod, water budget, and turnover time in
wetland studies can contribute to a better understanding of wetland function.
Hydrology affects species composition and richness, primary productivity,
organic accumulation, and nutrient cycling in wetlands.

The hydrology of a wetland creates the unique physiochemical conditions that make
such an ecosystem different from both well-drained terrestrial systems and deepwater
aquatic systems. Hydrologic pathways such as precipitation, surface runoff, ground-
water, tides, and flooding rivers transport energy and nutrients to and from wedands.
Water depth, flow patterns, and duration and frequency of flooding, which are the
result of all of the hydrologic inputs and outputs, influence the biochemistry of the
soils and are major factors in the ultimate selection of the biota of wetlands. Biota
ranging from microbial communities, to vegetation, to waterfowl are all constrained
or enhanced by hydrologic conditions. An important point about wetlands — one that

111

112 Chapter 4 Wetland Hydrology

is often missed by ecologists who begin to study these systems — is this: Hydrology is
probably the sinpile most important determinant of the establishment and maintenance
of specific types of wetlands and wetland processes. An understanding of rudimentary
hydrology should be in the repertoire of any wetland scientist.

Importance of Hydrology in Wetlands

Wetlands are transitional between terrestrial and open-water aquatic ecosystems.
They are transitional in terms of spatial arrangement, for they usually are found
between uplands and aquatic systems (see Fig. 2.1a). They are also transitional in
the amount of water they store and process and in other ecological processes that
result from the water regime. Wetlands form the aquatic boundary of the habitats
of many terrestrial plants and animals; they also form the terrestrial edge for many
aquatic plants and animals. Hence, small changes in hydrology can result in significant
biotic changes.

The starting point for the hydrolojjy of a wedand is the climate and basin geomor-
phology (Fig. 4.1). All things being equal, wetlands are more prevalent in cool or wet
climates than in hot or dry climates. Cool climates have less water loss from the land via
evapotranspiration, whereas wet climates have excess precipitation. The second impor-
tant factor is the geomorphology of the landscape and basin. Steep terrain tends to
have fewer wetlands than flat or gently sloping landscapes. Isolated basins have differ-
ent potential for wedands than do ddal-fed or river-fed environments. When climate,
basin geomorphology, and hydrology are considered as one unit, it is referred to as
a wedand’s hydrojjeomorpholojjy. Figure 4.1 illustrates that the hydrology of a wetland
directly modifies and changes its physiochemical environment (chemical and physical
properties), pardcularly oxygen availability and related chemistry, such as nutrient
availability, pH, and toxicity (e.g., the producdon of hydrogen sulfide). Hydrology
also transports sediments, nutrients, and even toxic materials into wedands, thereby
further influencing the physiochemical environment. Except in nutrient-poor wedands
such as bogs, water inputs are the major source of nutrients to wedands. Hydrology
also causes water outflows from wetlands that often remove biotic and abiodc material,
such as dissolved organic carbon, excessive salinity, toxins, and excess sediments and
detritus. Some modifications in the physicochemical environment, such as the buildup
of sediments, can modify the hydrology by changing the basin geometry or affecdng
the hydrologic inflows or outflows (pathway A in Fig. 4.1).

Modificadons of the physiochemical environment, in turn, have a direct impact on
the biota in the wedand. When hydrologic condidons in wedands change even slighdy,
the biota may respond with massive changes in species composidon and richness and
in ecosystem producdvity. Biota such as emergent aquadc plants adapt to the anoxia in
the sediments, although the anoxia excludes most vascular plant species. The level of
nutrients in the sediments determines producdvity and which species will dominate.
Animals adapted to shallow water and this vegetadon cover will flourish. Microbes able
to metabolize in anoxic condidons dominate the reduced sediments, while aerobic
microorganisms survive in a thin layer of oxidized sediments and in the water column

Importance of Hydrology in Wetlands 113

1. Hydrology

water

level.

flow.

frequency.

etc.

time

mcxiifies and determines

2. Physiochemical
environment

sediments, soil chemistry, water
chemistry, etc.

oxidized

reduced

allowing spedfic

3. Biota

vegetation, animals,
and microbes

which modifies

which, in turn,
modify the
physiochemistry
and

hydrology

■b

direct effect
-W. biotic feedback

Figure 4.1 Conceptual diagram illustrating the effects of hydrology on wetland function and
the biotic feedbacks that affect wetland hydrology. Pathways A and B are feedbacks to the
hydrology and physiochemistry of the wetland.

if oxygen is present there. When hydrologic patterns remain similar from year to year,
a wetland’s biotic structural and functional integrity may persist tor many years.

Biotic Controi of Wetiand Hydroiogy

Just as many other ecosystems exert feedback (cybernetic) control of their physical
environments, wedand biota are not passive to their hydrologic conditions. Pathway B
in Figure 4. 1 shows that the biotic components of wetlands can control the hydrology

114 Chapter 4 Wetland Hydrology

and chemistry of their environment through a variety of mechanisms. Microbes, in
particular, catalyze virtually all chemical changes in wetland soils and thus control
nutrient availability to plants and even the production of phytotoxins, such as sulfides.
Plants, animals, and microbes that use these essential biological feedback mechanisms
were formally recognized in the ecological literature as ecosystem engineers. Plants
cause changes in their physical environment through processes such as peat build-
ing, sediment trapping, nutrient retention, water shading, and transpiration. Wetland
vegetation influences the hydrologic conditions of the physicochemical environment
by binding sediments to reduce erosion, by trapping sediments, by interrupting water
flows, and by building peat. Accumulated sediments and organic matter, in turn, inter-
rupt water flows and can eventually decrease the duration and frequency by which
the wedands are flooded. Bogs build peat to the point at which they are no longer
influenced at the surface by the inflow of mineral waters. Some trees in some south-
ern swamps save water by their deciduous nature, their seasonal shading, and their
relatively slow rates of transpiration. In more temperate climates, trees that invade
shallow marshes and vernal pools can decrease water levels during the growing sea-
son by increasing transpiration, thus allowing even more woody plants to take over.
Removal of these trees in what appears to be a dry forest sometimes surprisingly causes
standing water and marsh vegetation to reappear.

Several animals are particularly noted for their contributions to hydrologic modifi-
cations and subsequent changes in wetlands. The exploits of beavers ( Castor canaden-
sis) in much of North America in both creating and destroying wetland habitats are
well known. They build dams on streams, backing up water across great expanses and
creating wedands where none existed before. In colonial times, beaver populations
covered the entire American continent north of Mexico, before fur trappers drasti-
cally reduced them. Beavers have been an important causal force in the creation of the
Great Dismal Swamp of Virginia and North Carolina. Hey and Philippi (1995) esti-
mated that a population of 40 million beavers could have accounted for 207,000 km^
of beaver ponds (wetlands) in the upper Mississippi and Missouri River basins before
European trappers entered the region and that, with the demise of the beaver, only I
percent of those beaver ponds exist today.

Muskrats ( Ondatra zibethicus) burrow through wetlands, changing flow patterns
and sometimes water levels directly. They harvest large amounts of emergent vege-
tation for their food and to build winter lodges, thereby opening up large areas of
marshes. Geese, especially Canada geese {Branta canadensis) and several varieties of
snow geese {Chen spp.), cause eat-outs^ or major wetland vegetation removal by her-
bivory, in many parts of the world. Newly planted wetlands are particularly susceptible
to Canada geese eat-outs in North America. By removing vegetation cover, these her-
bivores reset the successional status of the wetlands and thus have a major impact on
wetland hydrology.

The American alligator {Alligator mississippiensis) is known for its role in the
Florida Everglades in constructing “gator holes” that serve as oases for fish, turtles,
snails, and other aquatic animals during the dry season. In all of these cases, the biota of
the ecosystem have contributed to their own survival, to the survival of other species.

Wetland Hydroperiod 115

and to the elimination of others by mfluencing the ecosystem’s hydrology and other
physical characteristics.

Wetland Hydroperiod

The hydroperiod is the seasonal pattern of the water level of a wetland and is the wet-
land’s hydrologic signature. It characterizes each type of wetland, and the constancy
of its pattern from year to year ensures a reasonable stability for that wetland. It defines
the rise and fall of a wedand’s surface and subsurface water by integrating all of the
inflows and outflows. The hydroperiod is also influenced by physical features of the
terrain and by proximity to other bodies of water.

Many terms are used to describe qualitatively a wetland’s hydroperiod (Table 4.1).
These terms such as seasonally flooded or intermittently flooded are specific in their
meaning and should be used with care and with sufficient data in describing a wetland’s
hydroperiod. For wetlands that are not subtidal or permanently flooded, the amount
of time that a wetland is in standing water is called the flood duration. The average
number of times that a wetland is flooded in a given period is known as the flood
frequency. Both terms are used to describe periodically flooded wetiands such as coastal
salt marshes and riparian wedands.

Typical hydroperiods for a diverse set of wetlands are shown in Figure 4.2. A
coastal salt marsh has a hydroperiod of semidiurnal flooding and dewatering superim-
posed on a twice-monthly pattern of spring and ebb tides (Fig. 4.2a). Wetlands along
coastlines often show some of this same spring-and-ebb pulsing (Fig. 4.2b), whereas
others reflect seasonal water-level changes of freshwater inflows and the water levels

Table 4.1 Definitions of wetland hydroperiods

Tidal Wetlands

Subtidal — permanently flooded with tidal water

Irregularly exposed — surface exposed by tides less often than daily

Regularly flooded — alternately flooded and exposed at least once daily

Irregularly flooded — flooded less often than daily

Nontidal Wetlands

Permanently flooded — flooded throughout the year in all years

Intermittently exposed — flooded throughout the year except in years of extreme drought

Semipermanently flooded — flooded during the growing season in most years

Seasonally flooded — flooded for extended periods during the growing season but usually no surface
water by end of growing season

Saturated — substrate is saturated for extended periods during the growing season, but standing
water is rarely present

Temporarily flooded — flooded for brief periods during the growing season, but water table is
otherwise well below surface

Intermittently flooded — surface is usually exposed with surface water present for variable periods
without detectable seasonal pattern

Source: After Cowardin et al. (1979)

116 Chapter 4 Wetland Hydrology

b. t
0)
■a
(1)

Tidal salt marsh

yVVWM

MMi

V4/ViMf

Irregularly flooded ’’tidal'' swamp or marsh

Gulf coast freshwater marsh

Month

Figure 4.2 Hydroperiods for several different wetlands, presented in approximately the
same relative scale: (a) tidal salt marsh, Rhode Island; (b) irregularly flooded “tidal” swamp
or marsh; (c) Gulf Coast freshwater marsh, Louisiana; (d) Great Lakes marshes, northern
Ohio (natural and managed); (e) prairie pothoie marsh with little groundwater flow (dry and
wet years); (f) groundwater-fed prairie pothole marsh; (g) vernal pool, California; (h) sub-
tropical cypress dome, Florida; (i) aiiuviai swamp. North Carolina; (j) bottomiand hardwood
forest, northern Illinois; (k) mineral soil swamp, Ontario, Canada; (i) rich fen. North Waies;
(m) pocosin or Caroiina Bay, North Carolina; (n) tropical floodplain forest, Amazon River, Man-
aus, Brazil. (Data from Nixon and Oviatt, 1973; Mitsch et al., 1979; Gilman, 1982; Junk, 1982;
P.H. Zedler, 1987; Mitsch, 1989; van dor Vaik, 1989; Brinson, 1993; Woo and Winter, 1993)

of the ocean itself (Fig. 4.2c). Hydroperiods of coastal lacustrine wetlands along the
Laurentian Great Lakes in the United States and Canada vary considerably, depending
on whether pumps and water management are used or whether the marshes are open
to the seasonal patterns of river flows and lake levels (Fig. 4. 2d). In fact, the hydrope-
riods of these wetlands, when used as hunting clubs for waterfowl production, actually
are managed to be dry when the normal season calls for wet and wet when the seasonal

Wetland Hydroperiod 117

pattern calls for dry conditions. Water levels for interior wetlands, such as the prairie
potholes of North America, vary considerably from year to year (see the next section),
with differences depending on climate variability (Fig. 4.2e). Wetlands affected by
groundwater tend to have water levels that are less seasonally variable (Fig. 4.2f).

Some of the most seasonally variable wedands are the vernal pools of central Cal-
ifornia, where surface water essentially disappears in this Mediterranean-type climate
tor all but four or five months (Fig. 4.2g). Cypress domes in central Florida have
standing water during the wet summer season and dry periods in the late autumn and
early spring (Fig. 4.2h). Low-order riverine wetlands, such as the alluvial swamps in
the southeastern United States, respond sharply to local rainfall events rather than to

118 Chapter 4 Wetland Hydrology

•c

Figure 4.2 {Continued)

general seasonal patterns (Fig. 4.2i). The hydroperiods of many bottomland hardwood
forests and swamps in colder climates have distinct periods of surface flooding in the
winter and early spring due to snow and ice conditions followed by spring floods but
otherwise have a water table that can be a meter or more below the surface (Fig. 4.2j
and k).

Peatlands in cooler climates can have hydroperiods with little pronounced seasonal
fluctuation, as in the fen from North Wales (Fig. 4.21) . If peatlands such as the pocosins
of North Carolina are located in regions of warm summers, significant patterns of
seasonal water-level change will occur (Fig. 4.2m). The most dramatic hydroperi-
ods result from high-order rivers that are more influenced by seasonal patterns of
precipitation throughout a large watershed than by local precipitation, leading to a

Wetland Water Budget 119

more predictable and seasonally distinct hydroperiod. For example, the annual fluctu-
ation of water in the tropical floodplain forests along the Amazon River is a predictable
seasonal pattern that can include a seasonal fluctuation in water level of 5 to 10 m
caused by flooding of upstream rivers (Fig. 4.2n).

Year-to-Year Fluctuations

The hydroperiod is not the same each year but varies according to climate and
antecedent conditions. Great variability can be seen from year to year for some
wetlands, as illustrated in Figure 4.3 for a prairie pothole regional wetland in Canada
and for the Big Cypress Swamp region of south Florida. In the pothole region, a
wet-dry cycle of 10 to 20 years is seen; spring is almost always wetter than fall, but
depths vary significantly from year to year (Fig. 4.3a). Figure 4.3b illustrates cases
of an even seasonal rainfall pattern for the Big Cypress Swamp in Florida between a
fairly stable hydroperiod and a year with a significant dry season, which caused the
hydroperiod to vary about 1.5 m between high and low water. A three-year study
of groundwater levels in a red maple swamp shows dramatically different growing
season water levels from year to year (Fig. 4.4). Water is near or at the surface during
high precipitation periods (last half of first year and entire second year) while dry
low-water conditions are mainly driven by seasonal evapotranspiration in the swamp
accelerated by groundwater loss during tree transpiration.

Pulsing Water Levels

Water levels in most wetlands are generally not stable but fluctuate seasonally (ripar-
ian wetlands), daily or semidaily (types of tidal wedands), or unpredictably (wedands
in low-order streams and coastal wedands with wind-driven tides). Flooding “pulses”
that occur seasonally or periodically especially in riverine wedands nourish the wedands
with addidonal nutrients and carry away detritus and waste products. Pulse-fed wet-
lands are often the most producdve wedands and are the most favorable for expordng
materials, energy, and biota to adjacent ecosystems. Despite this obvious fact, many
wetland managers, especially those who manage wetlands for waterfowl, often attempt
to control water levels by isolating formerly open wedands with levees meant to restrict
flooding. A seasonally fluctuating water level, then, is the rule, not the excepdon, in
most wedands.

Wetland Water Budget

The hydroperiod, or hydrologic state of a given wedand, can be summarized as being
a result of these three factors:

1 . The balance between the inflows and outflows of water

2. The surface contours of the landscape

3. Subsurface soil, geology, and groundwater conditions

1965 1970 1975 1980 1985

Year

Figure 4.3 YeaMo-year fluctuations in wetland water levels in two regions: (a) spring and
fall water depths for 25 years in shallow open-water wetlands in the prairie pothole region
of southwestern Saskatchewan, Canada; and (b) wet and dry year hydrographs for the Big
Cypress Swamp region of the Everglades, southwestern Florida, ((a) After Kantrud et al.,
1989 and Millar, 1971; (b) after Freiberger, 1972 and Duever, 1988)

120

Wetland Water Budget 121

1985 1986 1987 1988

Date

Figure 4.4 Relative water levels in two seasonally saturated red maple swamps in Rhode
Island, United States, for 1985 to 1987. Growing season precipitation amounts for 1985,
1986, and 1987 were 104, 76, and 59 cm, respectively. (After Golet et al., 1993)

Figure 4.5 Generalized water budget for a wetland with corresponding terms as in
Equation 4.1. P — precipitation; ET = evapotranspiration; I = interception; P„ - net precip-
itation; S, - surface inflow; S, - surface outflow; G,. — groundwater inflow; G^ - groundwater
outflow; T = tide or seiche; AV/At — change in storage per unit time.

The first condition defines the water bud£iet of the wetland, whereas the second
and the third define the capacity of the wetland to store water. The general bal-
ance between water storage and inflows and outflows, illustrated in Figure 4.5, is
expressed as

^ = P, + 5,+ G,-£T-5„-G„±r (4.1)

where

V = volume of water storage in wedands
AF/Ar = change in volume of water storage in wedand per unit dme, t
= net precipitation

5,- = surface inflows, including flooding streams

122 Chapter 4 Wetland Hydrology

Gj = groundwater inflows
ET = evapotranspiration
S„ = surface outflows
G„ = groundwater outflows
T = tidal inflow (+) or outflow (— )

The average water depth, d, at any one time, can further be described as

(4.2)

where

A = wetland surface area

Each of the terms in Equation 4.1 can be expressed in terms of depth per unit
time (e.g., cm/yr) or in terms of volume per unit time (e.g., m^/yr).

Examples of Water Budgets

Equation 4.1 and Figure 4.5 serves as useful summaries of the major hydrologic com-
ponents of any wetland water budget. Examples of hydrologic budgets for several
wetlands are illustrated in Figure 4.6. The terms in the equation vary in importance
according to the type of wedand observed; furthermore, not all terms in the hydro-
logic budget apply to all wedands (Table 4.2). There is a large variability in certain
flows, pardcularly in surface inflows and outflows, depending on the openness of the

Table 4.2 Major components of hydrologic budgets for wetlands

Component

Pattern

Wetlands Affected

Precipitation

Varies with climate, although many
regions have distinct wet and dry
seasons

All

Surface inflows

Seasonally, often matched with

Potentially all wetlands except

and outflows

precipitation pattern or spring thaw;
can be channelized as streamflow or
nonchannelized as runoff; includes
river flooding of alluvial wetlands

ombrotrophic bogs; riparian wetlands,
including bottomland hardwood forests
and other alluvial wetlands, are
particularly affected by river flooding

Groundwater

Less seasonal than surface inflows
and not always present

Potentially all wetlands except
ombrotrophic bogs and other perched
wetlands

Evapotranspiration

Seasonal with peaks in summer and
low rates in winter. Dependent on
meteorological, physical, and
biological conditions in wetlands

All

Tides

One to two tidal periods per day;
flooding frequency varies with
elevation

Tidal freshwater and salt marshes;
mangrove swamps

Wetland Water Budget 123

Figure 4.6 Annual water budgets for several wetlands. See Figure 4.5 for symbol def-
initions. All values are expressed in centimeters per year (cm/yr) except (b), which is
March-September only. (Data from Pride et al., 1966; Shjeflo, 1968; Mitsch, 1979; Hemond,
1980; Gilman, 1982; Twilley, 1982; Richardson, 1983; Mitsch and Reeder, 1992; Mitsch et al.,
2010)

wetlands. An alluvial cypress swamp in southern Illinois received a gross inflow of
floodwater from one flood that was more than 50 times the gross precipitation tor
the entire year (Fig. 4.6a). Even the net surface inflow from that flood (the water left
behind after the flooding river receded) was three times the precipitation input for the
entire year. Surface and groundwater inflows to a coastal Lake Erie marsh in northern

124 Chapter 4 Wetland Hydrology

Figure 4.6 {Continued)

Ohio were estimated to be almost 20 times the precipitation for a major part of a
drought year (Fig. 4.6b), and tides contributed 10 times the precipitation to a black
mangrove swamp in Florida (Fig. 4.6c).

In contrast to these inflow-dominated wetlands, surface inflow is approximately
equal to the precipitation inflow in the prairie pothole marshes of North Dakota

Wetland Water Budget 125

(Fig. 4.6d), considerably less than the precipitation for the Okefenokee Swamp in
Georgia (Fig. 4.6e) and a rich fen in North Wales (Fig. 4.6f), and essentially nonexis-
tent in the upland Green Swamp of central Florida (Fig. 4.6g), a bog in Massachusetts
(Fig. 4.6h), and a pocosin wetland of North Garolina (Fig. 4.6i). In most of these
examples, the change in storage is small or zero, indicating that the water level at
the end of the study period (usually an annual cycle) is close to where it was at the
beginning of the study period.

The water budget for the tropical Okavango Delta in southern Africa (Botswana)
has been investigated for many years. Figure 4.6j represents the average conditions
for the past 36 years. The data show that the Okavango River input, when averaged
over the entire delta, is about equivalent to the rainfall over this vast area. Furthermore,
the budget shows that essentially all of the inputs are balanced by a loss of evapotran-
spiration in this semiarid climate, and only about 1 percent of the water now leaves
the wetland region to the downstream village of Maun.

Residence Time — How Long Does Water Stay in a Wetiand?

A generally useful concept of wetland hydrology is that of the renewal rate or
turnover rate of water, defined as the ratio of throughput to average volume
within the system:

Few measurements of renewal rates have been made in wetlands, although
the renewal rate is a frequently used parameter in limnological studies.
Chemical and biotic properties are often determined by the openness of the
system, and the renewal rate is an index of this because it indicates how
rapidly the water in the system is replaced. The reciprocal of the renewal
rate is the turnover time or residence time (t, sometimes called detention
time by engineers for constructed wetlands), which is a measure of the
average time that water remains in the wetland. The theoretical residence
time, as calculated as the reciprocal of Equation 4.3, is often much longer
than the actual residence time of water flowing through a wetland, because
of nonuniform mixing. Because there are often parts of wetland where
waters are stagnant and not well mixed, the theoretical residence time (t)
estimate should be used with caution when estimating the hydrodynamics of
wetlands.

(4.3)

where

= renewal rate (time~^)

Qt = total inflow rate (volume/time)

V = average volume of water storage in wetland

126 Chapter 4 Wetland Hydrology

Precipitation

Wetlands occur most extensively in regions where precipitation, a term that includes
rainfall and snowfall, is in excess of losses such as evapotranspiration and surface runoff.
The fate of precipitation that falls on a wetland with forested, shrub, or emergent vege-
tation is shown in Figure 4.7. When some of the precipitation is retained by the vegeta-
tion cover, particularly in forested wetlands, the amount that actually passes through

Figure 4.7 Fate of precipitation in (a) a forested wetland and (b) a marsh. P = precipitation;
TF = throughfall; SF — stemflow.

Surface Flow 127

the vegetation to the water or substrate below is called throu0hfall. The amount of
precipitation that is retained in the overlying vegetation canopy is called interception.
Interception depends on several factors, such as the total amount of precipitation, the
intensity of the precipitation, and the character of the vegetation, including the stage
of vegetation development, the type of vegetation (e.g., deciduous or evergreen), and
the strata of the vegetation (e.g., tree, shrub, or emergent macrophyte). The percent-
age of precipitation that is intercepted in forests varies between 8 and 35 percent. The
water budget in Figure 4.6a, for example, illustrates that 29 percent of precipitation
in a forested wetland was intercepted by a canopy dominated by the deciduous conifer
Taxodium distichum.

Littie is known about the interception of precipitation by emergent herbaceous
macrophytes, but it probably is similar to that measured in grasslands or croplands.
Essentially, in those systems, interception at maximum growth can be as high as that
in a forest (10 to 35 percent of gross precipitation). An interesting hypothesis about
interception and the subsequent evaporation of water from leaf surfaces is that, because
the same amount of energy is required whether water evaporates from the surface of
a leaf or is transpired by the plant, the evaporation of intercepted water is not “lost”
because it may reduce the amount of transpiration loss that occurs. This suggests that
wetlands with either high or low interception may have similar overall water loss to
the atmosphere.

Another term related to precipitation, stemflow, refers to water that passes down
the stems of the vegetation (Fig. 4.7). This flow is generally a minor component of
the water budget of a wetland. For example, Heimburg (1984) found that stemflow
was, at maximum, 3 percent of throughfall in cypress dome wetlands in north-central
Florida.

These terms are related in a simple water balance as follows:

P=I+TF+SF (4.4)

where

F = total precipitation
I = interception
TF = throughfall
SF = stemflow

The total amount of precipitation that actually reaches the water’s surface or sub-
strate of a wetland is called the net precipitation (F^) and is defined as

Pn = P-I (4.5)

Surface Flow

Watersheds and Runoff

The percentage of precipitation that becomes surface flow depends on several
variables, with climate being the most important. Humid cool regions such as

128 Chapter 4 Wetland Hydrology

the Pacific Northwest, western British Columbia, and the northeastern Canadian
provinces have 60 to 80 percent of precipitation converted to runoff. In the arid
southwestern United States, less than 10 percent of the already low precipitation
becomes runoff This difference is related, in large part, to the higher temperatures in
the arid Southwest, which translate into higher evapotranspiration rates, greater soil
moisture deficits, and higher soil infiltration rates than in the Pacific Northwest. Even
though runoff in arid regions is small relative to that in humid areas, it does contribute
streamflow^ which is an important part of a riparian wetland’s water budget. Wedands
can be receiving systems for surface water flows {inflows)^ or surface water streams
can originate in wetlands to feed downstream systems {outflows). Surface outflows
are found in many wedands that are located in the upstream reaches of a watershed.
These wedands are often important water flow regulators for downstream rivers.
Some wedands have surface outflows that develop only when their water stages
exceed a critical level.

Wetlands are subjected to surface inflows of several types. Overland flow is non-
channelized sheet flow that usually occurs during and immediately following rainfall or
a spring thaw, or as tides rise in coastal wetlands. A wedand influenced by a drainage
basin may receive channelized streamflow during most or all of the year. Wedands
are often an integrated part of a stream or river, for example, as instream freshwater
marshes or riparian bottomland forests. Wetlands that form in wide, shallow expanses
of river channels or floodplains adjacent to them are greatly influenced by the seasonal
streamflow patterns of the river. Wedands can also receive surface inflow from seasonal
or episodic pulses of flood flow from adjacent streams and rivers that may otherwise
not be connected hydrologically with the wedand. Coastal saline and brackish wedands
are also significantly influenced by freshwater runoff and streamflow (in addition to
ddes) that contribute nutrients and energy to the wedand and often ameliorate the
effects of soil salinity and anoxia.

Surface inflow from a drainage basin into a wedand is usually difficult to esdmate
without a great deal of data. Nevertheless, it is often one of the most important sources
of water in a wedand’s hydrologic budget. The direct runoff component of streamflow
refers to rainfall during a storm that causes an immediate increase in streamflow. An
esdmate of the amount of precipitadon that results in direct runoff, or quickflow., from
an individual storm can be determined from the following equadon:

S, = Rj,PA,, (4.6)

where

S^ = direct surface runoff into wedand (m^ per storm event)

Rp = hydrologic response coefficient
P = average precipitadon in watershed (m)

= area of watershed draining into wedand (m^)

This equadon states that the flow is propordonal to the volume of precipitadon
(Px A,,,) on the watershed feeding the wedand in quesdon. Rp, which represents the

Surface Flow 129

fraction of precipitation in the watershed that becomes direct surface runoff, ranges
from 4 to 18 percent for small watersheds in the eastern North America and generally
increases with latitude. Slope and type of vegetation appear to have little influence on
Rp in a watershed with a mature forest cover. As the following paragraph suggests,
land use and soil type can strongly influence runoff.

While Equation 4.6 predicts the volume of direct runoff caused by a storm event,
in some cases wetland scientists and managers might be interested in calculating the
peak runoff {flood peak) into a wetland caused by a specific rainfall event. Although
this is generally a difficult calculation for large watersheds, a formula with the unlikely
name of the rational runoff method is a widely accepted and useful way to predict peak
runoff for watersheds less than 80 ha in size. The equation is given by

= 0.278 CIA. (4.7)

where

^i{pk) — (pk) runoff into wetland (m^/s)

C = rational runoff coefficient (see Table 4.3)

I = rainfall intensity (mm/h)

= area of watershed draining into wetland (km^)

The coefficient C, which ranges from 0 to 1 (Table 4.3), depends on the upstream
land use. Concentrated urban areas have a coefficient ranging from 0.5 to 0.95, and

Table 4.3 Values of the rational runoff coefficient C used to calculate peak runoff

Urban Areas

Business areas:

high-value districts

0.75-0.95

neighborhood districts

0.50-0.70

Residential areas:

single-family dwellings

0.30-0.50

multiple-family dwellings

0.40-0.75

suburban

0.25-0.40

Industrial areas:

light

0.50-0.80

heavy

0.60-0.90

Parks and cemeteries

0.10-0.25

Playgrounds

0.20-0.35

Unimproved land

0.10-0.30

Rural Areas

Sandy and gravelly soils:

cultivated

0.20

pasture

0.15

woodland

0.10

Loams and similar soils:

cultivated

0.40

pasture

0.35

woodland

0.30

Heavy clay soils; shallow soils over bedrock:

cultivated

0.50

pasture

0.45

woodland

0.40

130 Chapter 4 Wetland Hydrology

rural areas have lower coefficients that greatly depend on soil type, with sandy soils
lowest (C = 0.1— 0.2) and clay soils highest (C = 0.4— 0.5).

Channelized Streamflow

Channelized streamflow into and out of wetlands is described simply as the product
of the cross-sectional area of the stream (A^) and the average velocity (v) and can be
determined through stream velocity measurements in the held:

Si or S„ = A^v (4.8)

where

S„ = surface channelized flow into or out of wedand (m^/s)

A^ = cross-sectional area of stream (m^)

V = average velocity (m/s)

The velocity can be determined in several ways, ranging from handheld veloc-
ity meter readings taken at various locations in the stream cross-section to the
floating-orange technique where the velocity of a floating orange or similar fruit
(which is 90 percent or more water and therelbre floats but just beneath the water
surface) is timed as it goes downstream. If a continuous or daily record of streamflow
is needed, then a rating curve (Fig. 4.8), a plot of instantaneous streamflow (as
estimated using Equation 4.8) versus stream elevation or stage, is useful. If this type
of rating curve is developed for a stream (the basis of most hydrologic streamflow
gauging stations operated by the U.S. Geological Survey), then a simple measurement
of the stage in the stream can be used to determine the streamflow. Because hydro-
graphs generally assume a constant water gradient, caution should be taken in using
this approach for streams flowing into wedands to ensure that no “backwater effect”
of the wedand’s water level will affect the stream stage at the point of measurement.

100 500 1000 5000 1000C

stream discharge, cubic feet per second (cfs)

Figure 4.8 Rating curve for streamfiow determination as a function of stream stage.
100 cfs - 2.832 m^/s. (After Dunne and Leopold, 1978)

Surface Flow 131

Measuring Streamflow with Weirs

When a weir or other control structure is used at the outflow of a wetland
(Fig. 4.9), the outflow of a wetland can be estimated to be a function of the
water level in the wetland itself according to the equation:

So = xLy (4.9)

where

Sq = surface outflow

L = wetland water level above a control structure crest
(level at which flow just begins)

X, y = calibration coefficients

Figure 4.9 Control structures such as the V-notched weir shown here can be used for
measuring surface water flow in small streams into or out of wetlands. (Photo by W. J.
Mitsch)

132 Chapter 4 Wetland Hydrology

If a control structure such as a rectangular or V-notched weir is used
to measure the outflow from a wetland, standard equations of the form of
Equation 4.9 can be obtained from water measurement manuals (e.g., U.S.
Department of Interior, 2001). Care should be taken to calibrate standard weir
equations with actual measurements of streamflow and water level.

When an estimate of surface flow into or out of a riverine wetland is needed and
no stream velocity measurements are available, the Manning equation often can be
used if the slope of the stream and a description of the surface roughness are known:

where

Si or

(4.10)

n = roughness coefficient (Manning coefficient; see Table 4.4)

R = hydraulic radius (m) (cross-sectional area divided by the wetted
perimeter; this is an estimate of the relative portion of the stream
cross section and hence flow volume, in contact with the streambed)
s = channel slope (dimensionless)

The equation states that flow is proportional to stream cross-section, as modi-
fied by the roughness of the streambed and the proportion of flow in contact with
that bed. Although the potential exists for their use in wetland studies, the roughness
coefficients given in Table 4.4 and the Manning equation (Eq. 4.10) have not been
used very often. The relationship is particularly useful for estimating streamflow where
velocities are too slow to measure directly and to estimate flood peaks from high-water
marks on ungauged streams. These circumstances are common in wetland studies.

Floods and Riparian Wetiands

A special case of surface flow occurs in wetlands that are in floodplains adjacent to rivers
or streams and are occasionally flooded by those rivers or streams. These ecosystems

Table 4.4 Roughness coefficients (n) for Manning equation used to
determine streamflow in natural streams and channels

Stream Conditions

Manning Coefficient, n

Straightened earth canals 0.02

Winding natural streams with some plant growth 0.035

Mountain streams with rocky streambed 0.040-0.050

Winding natural streams with high plant growth 0.042-0.052

Sluggish streams with high plant growth 0.065

Very sluggish streams with high plant growth 0.112

Surface Flow 133

1978 1979

j::

<13

Figure 4.10 River hydregraph from northeastern Iliinois, showing discharge and sediment
load of the river and discharge at which a riparian wetland is flooded (bankfull discharge).
1, 000 cfs = 28.32 m^/s. (After Bhowmik et al., 1980)

are often called riparian wetlands. The flooding of these wetlands varies in intensity,
duration, and number of floods from year to year, although the probability of flood-
ing is fairly predictable. In the eastern and midwestern United States and in much of
Canada, a pattern of winter or spring flooding caused by rains and sudden snowmelt is
often observed. When river flow begins to overflow onto the floodplain, the stream-
flow is referred to as bankfull discharge. A hydrograph of a stream that flooded its
riparian wetlands above bankfull discharge for several months in the spring is shown
in Figure 4.10. There is a remarkable consistency in the hydrographs of rivers in the
midwestern United States, in that they tend to overflow their banks (bankfull dis-
charge) at intervals between one and two years or on the average two years out of
three (see the next box).

Recurrence Interval

The recurrence interval is the average interval between the occurrences of
floods at a given or greater stage (depth). The inverse of the recurrence interval
is the average probability of flooding in any one year. Figure 4.11 suggests
that streams in the midwestern and southern United States will overflow their
banks onto the adjacent riparian forest with an average recurrence interval of
1.5 years (or a probability of 1/1.5, or 67 percent, of overbank flooding in any
one year). Stated another way, these rivers, on average, overflow their banks
in two out of every three years. Figure 4.11 also illustrates that flow that is
twice that of bankfull discharge occurs at recurrence intervals of approximately
five years; this flow, however, results in only a 40 percent greater river depth

134 Chapter 4 Wetland Hydrology

over bankfull depth on the floodplain. This predictable relationship suggests
that in natural stream systems, the size of a stream channel is related to the
hydraulic energy that scours the streambed.

Figure 4.11 Relationships among streamflow (discharge), stream depth, and recur-
rence interval for streams and rivers in the midwestern and southern United States. Q
— stream discharge; Q^f = bankfull discharge; d = stream depth; d^f = bankfull depth
(depth of river with floodplain is initially flooded). (After Leopold et al., 1964)

Groundwater

Recharge and Discharge Wetlands

Groundwater can heavily influence some wetlands, whereas in others it may have
hardly any effect at all. The influence of wetland recharge and discharge on groundwa-
ter resources has often been cited as one of the most important attributes of wedands,
but it does not hold for all wedand types; nor is there sufficient experience with
site-specific studies to make many generalizations. Groundwater inflow results when
the surface water (or groundwater) level of a wedand is lower hydrologically than the
water table of the surrounding land (called a discharjje wetland by geologists, who gen-
erally view their water budget from a groundwater, not from a wedand, perspecrive).
Wedands can intercept the water table in such a way that they have only inflows and

Groundwater 135

Figure 4.12 Possible discharge-recharge interchanges between wetlands and groundwater
systems including (a) marsh as a depression receiving groundwater flow (discharge wet-
land); (b) groundwater spring or seep wetland or groundwater slope wetland at the base of
a steep slope; (c) floodplain wetland fed by groundwater; (d) marsh as a recharge wetland
adding water to groundwater; (e) perched wetland or surface water depression wetland; (f)
groundwater flow through a tidal wetland. Dashed lines indicate groundwater level.

no outflows, as shown for a prairie marsh in Figure 4.12a. Another type of discharge
wetland, called a spring ot wetland, is often found at the base of steep slopes where
the groundwater surface intersects the land surface (Fig. 4.12b). This type of wetland
can be an isolated low point in the landscape; more often, it discharges excess water
downstream as surface water or as groundwater, as shown in the riparian wetland in
Figure 4.12c.

When the water level in a wetland is higher than the water table of its surround-
ings, groundwater will flow out of the wedand (called a recharge wetland^ Fig. 4.12d).
When a wedand is well above the groundwater of the area, the wedand is referred to
as being perched (Fig. 4.12e). This type of wedand, also referred to as a surface water
depression wetland, loses water only through infiltradon into the ground and through
evapotranspiradon. Tidally influenced wedands often have significant groundwater
inflows that can influence soil salinity and keep the wedand soil wet even during low
ride (Fig. 4.12f).

A final type of wedand, one that is fairly common, is litde influenced by ground-
water inflows. Because wedands often occur where soils have poor permeability,
the major source of water can be restricted to surface water runoff, with losses
occurring through evapotranspiradon and other outflows. This type of wedand

136 Chapter 4 Wetland Hydrology

often has fluctuating hydroperiods and intermittent flooding (e.g., prairie potholes
[Fig. 4.12e] and vernal pools [Fig. 4.12d], with standing water dependent on
seasonal precipitation and surface inflows. If, however, such a wedand were to be
influenced by groundwater, its water level would be better buffered against dramatic
seasonal changes (see Fig. 4.12a, c).

Nomenclature for the four types of groundwater hydrologic settings for freshwa-
ter wetlands are illustrated in Figure 4.13 and summarized here:

1. Surface water depression wetland (Fig. 4.13a). This type of wetland is
dominated by surface runoff and precipitation, with little groundwater
outflow due to a layer of low-permeability soils. This is similar to the perched
wetland type described in Figure 4.12e, where the wetland is separated from
the water table by an unsaturated zone.

2. Surface water slope wetland (Fig. 4.13b). This type of wetland is generally
found in alluvial soil adjacent to a lake or stream and is fed, to some degree,
by precipitation and surface runoff but, more important, by overbank

a. Surface water depression wetland

’ limited recharge

water table usually possible
below wetland

c. Groundwater depression wetland

seasonal recharge possible
when water table drops
below wetland

b. Surface water slope wetland

limited recharge
water table usually possible
below wetland

d. Groundwater slope wetland

Figure 4.13 Novitski groundwater fiow patterns for wetlands: (a) surface water depression,
(b) surface water slope, (c) groundwater depression, and (d) groundwater slope. Dashed lines
indicate groundwater level. (After Golet et al., 1993)

Evapotranspiration 137

flooding from the adjacent stream, river, or lake. Hydroperiods of these
wetlands match the seasonal patterns of the adjacent bodies of water, with
relatively rapid wetting and drying. Some groundwater recharge is possible,
but that groundwater soon discharges back to the stream, river, or lake.

3. Groundwater depression wetland (Fig. 4.13c). This is the groundwater
discharge wetland described previously (Fig. 4.12a), where the wetland is in a
depression low enough to intercept the local groundwater table. These kinds
of wetlands can occur in coarse-textured glaciofluvial deposits, where the
interchange between groundwater and surface water is enhanced by relatively
coarse soil material. Water-level fluctuations in these types of wetlands are less
dramatic than fluctuations in surface flow wetlands because of the relative
stability of the groundwater levels.

4. Groundwater slope wetland (Fig. 4.13d). Wetlands often develop on slopes or
hillsides where groundwater discharges to the surface as springs and seeps.
Groundwater flow into these wedands can be continuous or seasonal,
depending on the local geohydrology and on the evapotranspiration rates of
the wetland and adjacent uplands.

Darcy’s Law

Darcy’s law, an equation familiar to groundwater hydrologists, often describes the flow
of groundwater into and out of a wetland. This law states that the flow of groundwater
is proportional to (1) the slope of the piezometric surface (the hydraulic gradient) and
( 2 ) the hydraulic conductivity, or permeability^ the capacity of the soil to conduct water
flow. In equation form, Darcy’s law is given as

G=kA^s (4.11)

where

G = flow rate of groundwater (volume per unit time)
k = hydraulic conductivity or permeability (length per unit time)

= groundwater cross-sectional area perpendicular to the direction of flow
s = hydraulic gradient (slope of water table or piezometric surface)

Despite the importance of groundwater flows in the budgets of many wetlands,
there is poor understanding of groundwater hydraulics in wetlands, particularly in
those that have organic soils. The hydraulic conductivity of both organic and inorganic
wetland soils is discussed in more detail in Chapter 5: “Wedand Soils.”

Evapotranspiration

The water that vaporizes from water or soil in a wedand {evaporation), together
with moisture that passes through vascular plants to the atmosphere {transpiration),
is called evapotranspiration. The meteorological factors that affect evaporadon and

138 Chapter 4 Wetland Hydrology

transpiration are similar as long as there is adequate moisture, a condition that almost
always exists in most wetlands. The rate of evapotranspiration is proportional to the
difference between the vapor pressure at the water surface (or at the leaf surface) and
the vapor pressure in the overlying air. This is described in a version of Dalton’s law.

E=cf(u)(e,,-e,) (4.12)

where

E = rate of evaporation
c = mass transfer coefficient
f(u) = function of wind speed, u

Cjj, = vapor pressure at surface, or saturation vapor pressure at wet surface
= vapor pressure in surrounding air

Evaporation and transpiration are enhanced by the same meteorological condi-
tions, such as solar radiation or surface temperature, that increase the value of the
vapor pressure at the evaporating surface and by factors such as decreased humidity
or increased wind speed that decrease the vapor pressure of the surrounding air. This
equation assumes an adequate supply of water for capillary movement in the soil or for
access by rooted plants. When the water supply is limited (not a frequent occurrence
in wetlands), evapotranspiration is limited as well. Transpiration can also be physiolog-
ically limited in plants through the closing of leaf stomata despite adequate moisture
during periods of stress such as anoxia.

Direct Measurement of Wetland Evapotranspiration

Several direct measurement techniques can be used in wetlands to deter-
mine evapotranspiration. The classical reference method is the measurement
of evaporation from a water-filled pan, usually by measuring the weight loss, by
measuring the volume required to replace lost water over a period of time, or
by measuring the drop in water level. This is generally considered a measure-
ment of potential evaporation, since the evaporating surface is saturated. The
method is tedious and the results often poorly correlated with actual evapo-
ration from vegetated surfaces, because the transpiration, unsaturated soils,
winds, and shading effects of the plant canopy all influence the rate, often in
unknown ways. However, pan evaporation provides a reference evaporation
rate for comparison with other techniques. Furthermore, because wetland
soils tend to be saturated most of the time, the pan method may be more
accurate for wetlands than for terrestrial environments.

Wetland evapotranspiration can also be estimated by measuring the
change in water level of the water in the wetland itself. This method, illustrated
in Figure 4.14, can be calculated as follows:

ET = Sy(24h ± s)

(4.13)

Evapotranspiration 139

where

ET = evapotranspiration (mm/day)

Sy = specific yield of aquifer (unitless)

= 1.0 for standing-water wetlands
<1.0 for groundwater wetlands

h = hourly rise in water level from midnight to 4:00 A.M. (mm/h)
s = net fall {+) or rise (-) of water table or water surface in one day

12 12 12

midnight noon midnight

Figure 4.14 Diurnal water fluctuation in some wetlands can be used to estimate evap-
otranspiration as in Equation 4.13.

The pattern assumes active “pumping" of water by vegetation during the
day and a constant rate of recharge equal to the midnight-to-4:00 A.M. rate.
This method also assumes that evapotranspiration is negligible around mid-
night and that the water table around this time approximates the daily mean.
The water level is usually at or near the root zone in many wetlands, a neces-
sary condition for this method to measure evapotranspiration accurately.

Empirical Estimates of Wetland Evapotranspiration

Thornthwaite Equation

Evapotranspiration can be determined with any number of empirical equations that use
easily measured meteorological variables. One of the most frequently used empirical

140 Chapter 4 Wetland Hydrology

equations for evapotranspiration from terrestrial ecosystems, which has been applied
with some success to wetlands, is the Thornthwaite equation for potential evapotran-
spiration:

£T, = 16(10 T,//)" (4.14)

where

ET^ = potential evapotranspiration for month i (mm/month)
Tj = mean monthly temperature (°C)

I = local heat index ^

!=1

a = (0.675 X - 77.1 X /^ + 17,920 x / + 492,390) x lO"'^

Penman Equation

A second empirical relationship that has had many applications in hydrologic and agri-
cultural studies but relatively few in wetlands is the Penman equation (Penman, 1948;
Chow, 1964). This equation, based on both Dalton’s law and the energy budget
approach, is given as

ET =

/AH + 0.27EA
A -10.27 J

(4.15)

where

ET = evapotranspiration (mm/day)

A = slope of curve of saturation vapor pressure versus mean air temperature
(mmHg/°C)

E[ = net radiation (cal/cm^-day)

= Rf(l - a) - Rb
R^ = total shortwave radiation
a = albedo of wetland surface
Rf^ = effective outgoing longwave radiation=/(T^)

E^ = term describing the contribution of mass transfer to evaporation
= 0.35 (0.5 + 0.0062Su){e,„ - ej
u = wind speed 2m above ground (km/day)

tjy = saturation vapor pressure of water surface at mean air temperature (mmHg)
= vapor pressure in surrounding air (mmHg)

The Penman equation was compared with the pan evaporation (multiplied by a
factor of 0.8) and other methods at natural enriched fens in Michigan and constructed
wetlands in Nevada. The Penman equation, like the Thornthwaite equation, gener-
ally underpredicted evapotranspiration from the humid Michigan wetland but agreed
within a few percentage points with other measurement techniques for the arid Nevada
wetlands.

Because of the many meteorological and biological factors that affect evapotran-
spiration, none of the many empirical relationships is entirely satisfactory for esti-
mating wetland evapotranspiration. Several comparisons of approaches to measuring

Evapotranspiration 141

evapotranspiration have been attempted (Lott and Hunt, 2001; Rosenberry et al.,
2004). One finding has been that empirical estimates of potential evapotranspiration
(PET), such as those determined from the Penman equation, generally underestimate
true wetland evapotranspiration during the growing season, possibly due to limitation
of the equation for describing surface roughness. A comparison of an energy bud-
get method for estimating evapotranspiration at a wetland in North Dakota with 12
empirical evapotranspiration equations found that most of the empirical methods gave
reasonable approximations of evapotranspiration (Rosenberry et al., 2004).

The Thornthwaite equation, the simplest method investigated as it only requires
air temperature, worked relatively well and may provide the most accurate measure-
ment per instrument cost. It remains one of the more commonly uses empirical
equations for estimating wetland evapotranspiration, but it only gives monthly
estimates, not daily or hourly rates.

Effects of Vegetation on Wetland Evapotranspiration

A question about evapotranspiration from wetlands that does not elicit a uniform
answer in the literature is: “Does the presence of wetland vegetation increase or
decrease the loss of water compared to that which would occur from an open body
of water.>” Data from individual studies are conflicting. Obviously, the presence of
vegetation retards evaporation from the water surface, but the question is whether
the transpiration of water through the plants equals or exceeds the difference. Eggels-
mann (1963) found evaporation from bogs in Germany to be generally less than that
from open water except during wet summer months. In studies of evapotranspiration
from small bogs in northern Minnesota, Bay (1967) found it to be 88 percent to
121 percent of open-water evaporation. Eisenlohr (1976) reported 10 percent lower
evapotranspiration from vegetated prairie potholes than from nonvegetated potholes
in North Dakota. Hall et al. (1972) estimated that a stand of vegetation in a small
New Hampshire wetland lost 80 percent more water than did the open water in the
wetland. In a forested pond cypress dome in north-central Florida, Heimburg (1984)
found that swamp evapotranspiration was about 80 percent of pan evaporation during
the dry season (spring and fall) and as low as 60 percent of pan evaporation during
the wet season (summer). S. L. Brown (1981) found that transpiration losses from
pond cypress wetlands were lower than evaporation from an open-water surface even
with adequate standing water.

In the arid West, it has been a long-standing practice to conserve water for irriga-
tion and other uses by clearing riparian vegetation from streams. In this environment
where groundwater is often well below the surface but within the rooting zone of
deep-rooted plants, trees “pump” water to the leaf surface and actively transpire even
when little evaporation occurs at the soil surface.

The conflicting measurements and the difficulty of measuring evaporation and
evapotranspiration led Linacre (1976) to conclude that neither the presence of wet-
land vegetation nor the type of vegetation had major influences on evaporation rates,
at least during the active growing season. Bernatowicz et al. (1976) also found lit-
tle difference in evapotranspiration among several species of vegetation. The general

142 Chapter 4 Wetland Hydrology

unimportance of plant species variation on overall wetland water loss is probably a
reasonable conclusion for most wetlands, although it is clear that the type of wetland
ecosystem and the season are important considerations. Ingram (1983), for example,
found that fens have about 40 percent more evapotranspiration than do treeless bogs
and that evaporation from the bogs is less than potential evapotranspiration in the
summer and greater than potential evapotranspiration in the winter.

In some cases, the type of vegetation in the wetiand does matter. When trees are
removed from some forested swamps where the soil is hydric but there is little surface
flooding, standing water may return and, with it, herbaceous marsh vegetation. This
resets a hydrologic succession; woody plants are able to reinvade the marsh during dry
years and reestablish the site back to a forested wetland.

Tides

The periodic and predictable tidal inundation of coastal salt marshes, mangroves, and
freshwater tidal marshes is a major hydrologic feature of these wetlands. The tide acts
as a stress by causing submergence, saline soils, and soil anaerobiosis; it acts as a subsidy
by removing excess salts, reestablishing aerobic conditions, and providing nutrients.
Tides also shift and alter the sediment patterns in coastal wetlands, causing a uniform
surface to develop.

Typical tidal patterns for several coastal areas of the United States are shown in
Figure 4.15a. Seasonal as well as diurnal patterns exist in the tidal rhythms. Annual
variations of mean monthly sea level are as great as 25 cm (Fig. 4.15b). Tides also
have significant bimonthly patterns, because they are generated by the gravitational
pull of the moon and, to a lesser extent, the sun. When the sun and the moon are
in line and pull together, which occurs almost every two weeks, spring tides^ or tides
of the greatest amplitude, develop. When the sun and the moon are at right angles,
neap tides, or tides of least amplitude, occur. Spring tides occur roughly at full and
new moons, whereas neap tides occur during the first and third quarters.

Tides vary more locally than regionally. The primary determinant is the coasdine
configuration. In North America, tidal amplitudes vary from less than I m along the
Texas Gulf Coast to several meters in the Bay of Fundy in Canada. Tidal amplitude can
actually increase as one progresses inland in some funnel-shaped estuaries. Typically,
on a rising tide, water flows up tidal creek channels until the channels are bankfull. It
overflows first at the upstream end, where tidal creeks break up into small creeks that
lack natural levees. The overflowing water spreads back downstream over the marsh
surface. On falling tides, the flows are reversed. At low tides, water continues to drain
through the natural levee sediments into adjacent creeks because these sediments tend
to be relatively coarse; in the marsh interior, where sediments are finer, drainage is poor
and water is often impounded in small depressions in the marsh.

Seiches

While inland wetlands are nontidal by definition, periodic water-level fluctuations
in wetlands adjacent to large freshwater lakes do occur as a result of short-term

Effects of Hydrology on Wetland Function 143

Figure 4.15 Patterns of tides: (a) daiiy tides for a month and (b) seasonai changes in mean
monthly sea levei for severai locations in North America. (After Emery and Uchupi, 1972)

water-level seiches, or “wind tides.” These are a common occurrence in wetlands
adjacent to large lakes, such as the Laurentian Great Lakes in the United States and
Canada (Fig. 4.16). When wind has a persistent direction, particularly in a long
fetch across a lake, water “piles up” on the downwind side of the lake, causing
high-water events for wetlands in that location. When the wind shifts or dies down,
the high water is released and flows to the opposite shoreline, causing a secondary
wind-relaxation seiche there and lower-than-normal water in the original high-water
location.

Effects of Hydrology on Wetland Function

The effects of hydrology on wetland structure and function can be described with a
complicated series of cause-and-effect relationships. A conceptual model of the general

144 Chapter 4 Wetland Hydrology

no wind

Figure 4.16 Concept of a seiche: a wind-relaxation seiche caused by (a) a steady wind that
(b) relaxes or shifts directions from initial wind set and (c) results in an oppositeiy directed
tilt; (d) water levels in Ohio (Toledo and Cleveland) and New York (Buffalo) coastlines of Lake
Erie during an Aprii 1979 storm and subsequent wind-reiaxation seiche. (After Korgen, 1995)

effects of hydrology in wetland ecosystems was shown in Figure 4.1. The effects are
shown to be primarily on the chemical and physical aspects of the wetlands, which, in
turn, influence the biotic components of the ecosystem. The biotic components then
have a feedback effect on hydrology. Four principles underscoring the importance of
hydrology in wetlands can be elucidated from studies that have been conducted to
date. These principles are described next.

Elevation above sea level,

Effects of Hydrology on Wetland Function 145

01 03 05 07 09 11 13 15 17 19 21 23 01 03 05 07 09 11 13 15 17 19 21 23 01 03 05 07 09 11 13 15 17

Aprils, 1979 Aprils, 1979

April 7, 1979

Figure 4.16 {Continued)

1. Hydrology leads to a unique vegetation eomposition but ean limit or enhance
species richness.

Hydrology is a two-edged sword for species composition and diversity in
wetlands. It acts as a limit or a stimulus to species richness, depending on
the hydroperiod and physical energies. At a minimum, the hydrology acts to

146 Chapter 4 Wetland Hydrology

select water-tolerant vegetation in both freshwater and saltwater conditions
and to exclude flood-intolerant species. Of the thousands of vascular plants on
Earth, relatively few have adapted to waterlogged soils. Although it is difficult
to generalize, many wetlands that sustain long flooding durations have lower
species richness in vegetation than do less frequently flooded or pulsing areas.
Waterlogged soils and the subsequent changes in oxygen content and other
chemical conditions significantly limit the number and the types of rooted plants
that can survive in this environment.

In general, species richness, at least in the vegetation community, increases
as flow-through or pulsing hydrology increases. Flowing water can be thought
of as a stimulus to diversity, probably caused by its ability to renew minerals and
reduce anaerobic conditions. Hydrology also stimulates diversity when the action
of water and transported sediments creates spatial heterogeneity, opening up addi-
tional ecological niches. When rivers flood riparian wetlands or when tides rise
and fall in coastal marshes, erosion, scouring, and sediment deposition sometimes
create niches that allow diverse habitats to develop. However, flowing water can
also create a relatively uniform surface that might allow monospecific stands of
Typha or FhmgmiUs to dominate a freshwater marsh or Spartina to dominate a
coastal marsh. Keddy (1992) likened water-level fluctuations in wetlands to fires
in forests. They eliminate one growth form of vegetation (e.g., woody plants) in
favor of another (e.g., herbaceous species) and allow regeneration of species from
buried seeds.

2. Primary productivity and other ecosystem functions in wetlands are often
enhanced by flowing conditions and a pulsing hydroperiod and are often
depressed by stagnant conditions.

In general, the “openness” of a wedand to hydrological fluxes is probably one of
the most important determinants of potential primary productivity. For example,
peatiands that have flow-through conditions (fens) have long been known to be
more productive than stagnant raised bogs. Some studies have found that wetlands
in stagnant (nonflowing) or continuously deep water have low productivities,
whereas wetlands that are in slowly flowing strands or are open to flooding rivers
have high productivities.

This relationship between hydrology and ecosystem primary productivity has
been investigated most extensively for forested wetlands. Figure 4.17 shows a
set of similar typical “Shelford-type” limitation curves that have been suggested
in separate studies to explain the importance of hydrology on forested wetland
productivity. All of the curves in Figure 4.17 suggest that the highest productivity
occurs in systems that are neither very wet nor too dry but that have either average
hydrologic conditions or seasonal hydrologic pulsing.

The subsidy-stress model of H. T. Odum (1971) and E. P. Odum (1979),
later refined as the pulse stability concept by all three Odums (W. E. Odum et ah,
1995 ), includes concepts that potentially apply well to the effects of hydrology on

flooding regime

Figure 4.17 Relationships between swamp productivity and hydrologic conditions:

(a) for cypress {Taxodium) swamps in north-central Florida, (b) between flooding regime
and net primary productivity of Louisiana swamps, and (c) between radial growth of
red maple (Acer rubrum) and annual water level for six Rhode Island red maple swamps
over six years, ((a) after Mitsch and Ewel, 1979; (b) after Conner and Day, 1982;

147

148 Chapter 4 Wetland Hydrology

Figure 4.17 {Continued)

wetiand productivity. Seasonal pulsing of floodwater can be both a subsidy and
a stress, whether the wetland is a salt marsh or mangrove swamp subject to
twice-per-day flooding or a riparian wetland subject to seasonal river pulses.
Pulsing is frequent in nature, and ecosystems such as bottomland forests and
salt marshes appear to be well adapted to taking advantage of this subsidy.
Despite this clear theoretical basis for understanding the effects of hydrology on
productivity, it has been difficult to confirm or deny these theories in practice.

The model shown m Figure 4.18 may explain the difficulty in ascribing a
direct relationship between vascular plant productivity and hydrologic conditions.
While flood intensity increases available moisture and nutrients, longer flood
durations increase stresses caused by an anaerobic root zone and can actually
decrease the length of the growing season. In effect, “subsidies and stresses
may occur simultaneously and cancel one another” (Megonigal et al., 1997). In
this Mitsch-Rust model, flood intensity and duration affect moisture, available
nutrients, anaerobiosis, and even length of growing season in a complex and
nonlinear “push-pull” arrangement.

The influence of hydrologic conditions on freshwater marsh productivity is
less certain. If peak biomass or similar measures are used as indicators of marsh

Effects of Hydrology on Wetland Function 149

Figure 4.18 Causal model that describes the major causes for increases and decreases in
individual tree growth in riparian floodplain forests. Plus (+) sign indicates a positive effect;
minus sign (— ) indicates a negative effect. (After Mitsch and Rust, 1984)

productivity, some studies have shown the classical stimulation of vegetation along
the water’s edge, whereas other studies have indicated a higher macrophyte pro-
ductivity in sheltered, nonflowing marshes than in wetlands that are open to
flowing conditions or coastal influences. For example, consistently higher macro-
phyte biomass was found in wetlands isolated from surface fluxes with artificial
dikes than in wetlands that were open to coastal fluxes along Lake Erie. Several
explanations are possible: (1) The coastal fluxes may also be serving as a stress as
well as a subsidy on the macrophytes; (2) the open marshes may be exporting a
significant amount of their productivity; and (3) the diked wetlands have more
predictable hydroperiods.

Similar results were found in a hydrologic pulsing experiment in central Ohio,
where simulated river floods caused a decrease in macrophyte and water column
primary productivity but led to changes in greenhouse gas emissions because of
a flushing effect (Mitsch et ah, 2005; Altor and Mitsch, 2006, 2008; Hernandez
and Mitsch, 2006, 2007; Tuttle et ah, 2008; Fig. 4.19a) Conversely, an earlier
study in Illinois of the influence of flow-through conditions on water column
primary productivity of constructed marshes found that, after two years of exper-
imentation, water column (phytoplankton and submerged aquatics) productivity
was higher in high-flow wetlands compared to low-flow wetlands (Fig. 4.19b).
While macrophyte productivity may take many years to respond to the difference

150 Chapter 4 Wetland Hydrology

pulsed hydrology steady-flow hydrology

0.8

Figure 4.19 Aquatic primary productivity in freshwater marshes as a function of hydroiogic
conditions: (a) puised flooding versus steady-flew hydrology at the Olentangy River Wetland
Research Park, central Ohio; (b) high-fiow and iow-fiow conditions at the Des Plaines Wet-
iand Demonstration Project, northeastern Illinois. * indicates statistical differences (0.05_
between low-flow and high-flow conditions, ((a) After Tbttle et al., 2008; (b) after Cronk and
Mitsch, 1994)

in hydrology, water column productivity, which is often caused by attached and
planktonic algae, responds relatively quickly to changing hydrologic conditions.

Coastal wetlands subject to frequent tidal action are generally more produc-
tive than those that are only occasionally inundated. A comparison of several
Atlantic Coast salt marshes, for example, showed a direct relationship between
tidal range (as a measure of water flux) and end-of-season peak biomass of
Spartina alterniflora (Fig. 4.20). Apparently, vigorous tides increase the nutrient
subsidy and cause a flushing of toxic materials, such as salt. Freshwater tidal

Effects of Hydrology on Wetland Function 151

Figure 4.20 Production of Spartina alternWora versus mean tidal range for severai Atlantic
Coast sait marshes. Different symbeis indicate different data sources. (After Steever
et al., 1976)

wetlands are even more productive than saline tidal wetlands, because they
receive the energy and nutrient subsidy of tidal flushing while avoiding the stress
of saline soils.

3. Accumulation of organic material in wetlands is controlled by hydrology
through its influence on primary productivity, decomposition, and export of
particulate organic matter.

Wetlands can accumulate excess organic matter as a result of either increased
primary productivity (as described previously) or decreased decomposition and
export. Notwithstanding the discrepancies from short-term litter decomposition
studies, peat accumulates to some degree in all wetlands as a result of these pro-
cesses. The effects of hydrology on decomposition pathways are even less clear
than the effects on primary productivity discussed previously. Probably the lack
of agreement among the many studies published on the subject results from the
complexity of the decomposition process. In general, decomposition of organic
detritus requires electron donors (usually oxygen, but alternate chemicals such as
sulfate or nitrate may be effective under anoxic conditions), moisture, inorganic

152 Chapter 4 Wetland Hydrology

nutrients, and microorganisms capable of metabolizing in the specific environ-
ment concerned. The observed rate of organic decomposition is also influenced
by the ambient temperature and by the activity of macrodetritivores that shred the
plant remains and/or repackage it as bacterially inoculated fecal pellets. Hydrol-
ogy modifies many of these variables; for example, moisture depends on the flood-
ing regime, flowing water carries oxygen and nutrients, while in stagnant water
oxygen is rapidly depleted and nutrients are transformed to more or less available
forms. Given this complexity, it is not surprising that the results of short-term in
situ decomposition studies often disagree.

The importance of hydrology for organic carbon export is obvious. A gen-
erally higher rate of export is to be expected from wetlands that are open to
the flowthrough of water. Riparian wetlands often contribute large amounts of
organic detritus to streams, including macrodetritus such as whole trees. There
is also considerable evidence that watersheds that drain wedand regions export
more organic material but retain more nutrients than do watersheds that do not
have wetlands (Fig. 4.21). For example, the slope of the line in Figure 4.21 for
wetland-dominated watersheds is much steeper than that for upland watersheds.

Figure 4.21 Organic carbon export from wetland-dominated watersheds cempared with non-
wetland watersheds. (From Mulholland and Kuenzier, 1979)

Techniques for Wetland Hydrology Studies 153

indicating a much greater organic carbon concentration in runoff as well as greater
export for a given runoff from the wetland -dominated watersheds. Salt marshes
and mangrove swamps are also considered major exporters of their productivity
by most, but the generality of this concept is not fully accepted by coastal ecol-
ogists. Hydrologically isolated wedands, such as northern peadands, have much
lower organic export.

4. Nutrient cycling and nutrient availability are both significantly infiuenced by

hydrologic conditions.

Nutrients are carried into wedands by the hydrologic inputs of precipitadon, river
flooding, ddes, and surface and groundwater inflows. Outflows of nutrients are
controlled primarily by the outflow of water. These hydrologic/nutrient flows
are also important determinants of wedand producdvity and decomposidon (see
previous sections). Intrasystem nutrient cycling is generally, in turn, ded to path-
ways such as primary productivity and decomposition. When productivity and
decomposition rates are high, as in flowing water or pulsing hydroperiod wet-
lands, nutrient cycling is rapid. When productivity and decomposidon processes
are slow, as in isolated ombrotrophic bogs, nutrient cycling is also slow.

The hydroperiod of a wedand has a significant effect on nutrient transfor-
madons, on the availability of nutrients to vegetation, and on loss from wetland
soils of nutrients that have gaseous forms. Thus, nitrogen availability and loss are
affected in wetlands by the reduced conditions that result from waterlogged soil.
Typically, a narrow oxidized surface layer develops over the anaerobic zone in wet-
land soils, causing a combination of reactions in the nitrogen cycle — nitrification
and denitriflcadon — that may result in substandal losses of dinitrogen gas to the
atmosphere. Furthermore, ammonium nitrogen is usually the form of nitrogen
most available to plants in wedand soils, because the anaerobic environment favors
the reduced ionic form over the nitrate common in agricultural soils.

Flooding of wedand soil, by altering both the pH and the redox potendal of
the soil, influences the availability of other nutrients. The pH of both acid and
alkaline soils tends to converge on a pH of 7 when they are flooded. The redox
potendal, a measure of the intensity oxidadon or reducdon of a chemical or bio-
logical system, indicates the state of oxidadon (and, hence, availability) of several
nutrients. Phosphorus is known to be more soluble under anaerobic condidons,
pardy because of the hydrolysis and reduction of ferric and aluminum phosphates
to more soluble compounds. The availability of major ions, such as potassium
and magnesium, and several trace nutrients, such as iron, manganese, and sulfur,
is also affected by hydrologic conditions in the wedands.

Techniques for Wetland Hydrology Studies

It is curious that so litde attendon has been paid to hydrologic measurements in
wedand studies, despite the importance of hydrology in ecosystem funcdon. A great

154 Chapter 4 Wetland Hydrology

Surface

flow

weir

Stream

gauge

Upland

Water level
recorder

piezometers

Pyrometer Windspeed/
(solar energy) direction

Throughfall

gauge

Precipitation
gauge

Evaporation
pan

** %

Staff \
gauge oo^ynstream
piezometers

Figure 4.22 Placement ef hydrology instruments in the landscape to estimate a water bud-
get for a floodplain wetland.

deal of information can be obtained with only a modest investment in supplies and
equipment. A diagram summarizing many of the hydrology measurements typical
for developing a wetland’s water budget is given in Figure 4.22. Water levels can be
recorded continuously with water-level recorders or data loggers or during site visits
with a staff gauge. With records of water level, all of the following hydrologic param-
eters can be determined: hydroperiod, frequency of flooding, duration of flooding,
and water depth. Water-level recorders can also be used to determine the change in
storage in a water budget, as in Equation 4.1.

Evapotranspiration measurements are more difficult to obtain, but several empir-
ical relationships, such as the Thornthwaite equation, use meteorological variables.
Evaporation pans can also be used to estimate total evapotranspiration from wet-
lands, although pan coefficients are highly variable. Evapotranspiration of continu-
ously flooded nontidal wetlands can also be determined by monitoring the diurnal
water-level fluctuation.

Precipitation or throughfall or both can be measured by placing a statistically
adequate number of rain gauges in random locations throughout the wetland or by
utilizing weather station data. Surface runoff to wetlands can usually be determined
as the increase in water level in the wetland during and immediately following a storm
after net precipitation has been subtracted. Weirs can be constructed on more perma-
nent streams to monitor surface water inputs and outputs.

Groundwater flows are usually the most difficult and most costly hydrologic flows
to measure accurately. In some cases, clusters of shallow monitoring wells, placed
around a wetland, will help indicate the direction of groundwater flow and the slope
of the water or hydraulic gradient as required in Equation 4.11. The wells are called
piezometers when they are only partially screened, and thus measure the piezometric
head of an isolated part of the groundwater rather than being screened through the

References 155

entire length of the well and thus measuring the surface water aquifer. Piezometers can
be installed by professional well-drilling companies or, for low-budget installations,
generally can be installed with augers or as well points. Estimates of permeability or
hydraulic conductivity are then required to quantify the flows. Permeability can be
estimated through in situ pump tests using the wells or through laboratory analysis
of intact soil cores. The variability of results among different hydraulic conductivity
measuring techniques suggests that caution should be used in taking these numbers.

If a wetland is a perched or a recharge wedand, seepage can be estimated either
through a water budget approach (e.g., subtracting evapotranspiration losses from
water-level decreases when there are no other inflows or outflows) or by using
half-barrel seepage meters. Other methods available to measure groundwater flows in
wetlands include the use of stable isotopes, generally or because of

the propensity of the lighter isotope in each case to evaporate more readily, allowing
water to be “tagged” according to its source (Hunt et ah, 1996). Groundwater flow
models have also been used to estimate the flow of groundwater into and out of
wetlands with some success (Hunt et ah, 1996; Koreny et al., 1999).

The uncertainty in the scientific literature concerning many wetland processes
(e.g., the rates of organic matter decomposition discussed earlier) is often closely
related to unquantified hydrologic parameters. Thus, careful attention to quantifica-
tion of pertinent hydrologic parameters in wedand research studies is virtually certain
to improve our understanding of the ecological processes that control wetlands.

Recommended Readings

Brooks, K. N., P. F. Ffolliott, and J. A. Magner. 2012. Hydrology and Management of
Watersheds, 4th ed. Chichester, UK: Wiley-Blackwell.

Winter, T. C., and M. R. Flamas, eds., 1993. Hydrojjeolojyy of Wetlands. Special Issue
of Journal of Hydrolojjy 141 :l-269.

References

Altor, A. E., and W. J. Mitsch. 2006. Methane flux from created riparian marshes:
Reladonship to intermittent versus condnuous inundadon and emergent macro-
phytes. Ecolopfical Engineering 2B\ 224-234.

Altor, A. E., and W. J. Mitsch. 2008. Pulsing hydrology, methane emissions, and car-
bon dioxide fluxes in created marshes: A 2-year ecosystem study. Wetlands 28:
423-438.

Bay, R. R. 1967. Groundwater and vegetadon in two peat bogs in northern Min-
nesota. Ecolojjy 48: 308-310.

Bernatowicz, S., S. Leszczynski, and S. Tyczynska. 1976. The influence of transpira-
don by emergent plants on the water balance of lakes. Aquatic Botany 2: 275-288.

156 Chapter 4 Wetland Hydrology

Bhowmik, N. G., A. P. Bonini, W. C. Bogner, and R. P. Byrne. 1980. Hydraulics of
flow and sediment transport to the Kankakee River in Illinois. Illinois State Water
Survey Report of Investigation 98, Champaign, IL. 170 pp.

Brinson, M. M. 1993. Changes in the functioning of wetlands along environmental
gradients. Wetlands 13: 65-74.

Brown, S. L. 1981. A comparison of the structure, primary productivity, and transpi-
ration of cypress ecosystems in Florida. Ecological Monographs 51: 403-427.

Chow, V. T., ed. 1964. Handbook of Applied Hydrology. McGraw-Hill, New York.
1453 pp.

Conner, W. H., and J. W. Day, Jr. 1982. The ecology of forested wetlands in the
southeastern United States. In B. Copal, R. E. Turner, R. G. Wetzel, and D. F.
Whigham, eds.. Wetlands: Ecology and Manapiement . National Institute of Ecol-
ogy and International Scientific Publications, Jaipur, India, pp. 69-87.

Cronk, J. K., and W. J. Mitsch. 1994. Aquatic metabolism in four newly constructed
freshwater wetlands with different hydrologic inputs. Ecological Enpiineerinp; 3:
449-468.

Duever, M. J. 1988. Hydrologic processes for models of freshwater wetlands. In W. J.
Mitsch, M. Straskraba, and S. E. Jorgensen, eds.. Wetland Modelling. Elsevier,
Amsterdam, The Netherlands, pp. 9-39.

Dunne, T., and L. B. Leopold. 1978. Water in Environmental Planning . W. H. Free-
man and Company, New York. 818 pp.

Eggelsmann, R. 1963. Die Potentielle und Aktuelle Evaporation eines Seeklimath-
ochmoores. Publication 62, International Association of Hydrological Sciences,
pp. 88-97.

Eisenlohr, W. S. 1976. Water loss from a natural pond through transpiration by
hydrophytes. Water Resources Research 2: 443^53.

Emery, K. O., and E. Uchupi. 1972. Western North Atlantic Ocean: Topography,
rocks, structure, water, life, and sediments. Memoirs of the American Association
of Petroleum Geologists 17. 1532 pp.

Freiberger, H. J. 1972. Streamflow Variation and Distribution in the Bijj Cypress
Watershed During Wet and Dry Periods. Map Series 45, Bureau of Geology,
Florida Department of Natural Resources, Tallahassee.

Gilman, K. 1982. Nature conservation in wedands: Two small fen basins in western
Britain. In D. O. Logofet and N. K. Luckyanov, eds.. Ecosystem Dynamics in Fresh-
water Wetlands and Shallow Water Bodies, Vol. I. SCOPE and UNEP Workshop,
Center of International Projects, Moscow, pp. 290-310.

Golet, F. C., A. J. K. Calhoun, W. R. DeRagon, D. J. Lowry, and A. J. Gold. 1993.
Ecolojjy of Red Maple Swamps in the Glaciated Northeast: A Community Profile.
Biological Report 12, U.S. Fish and Wildlife Service, Washington, DC. 151 pp.

References 157

Hall, F. R., R. J. Rutherford, and G. L. Byers. 1972. The Influence of a New England
Wetland on Water Quantity and Quality. New Hampshire Water Resource Center
Research Report 4. University of New Hampshire, Durham. 51 pp.

Heimburg, K. 1984. Hydrology of north-central Florida cypress domes. In K. C. Ewel
and H. T. Odum, eds.. Cypress Swamps . University Presses of Florida, Gainesville,
pp. 72-82.

Hernandez, M. E., and W. J. Mitsch. 2006. Influence of hydrologic pulses, flood-
ing frequency, and vegetation on nitrous oxide emissions from created riparian
marshes. Wetlands 26: 862-877.

Hernandez, M. E., and W. J. Mitsch. 2007. Denitrification in created riverine wet-
lands: Influence of hydrology and season. Ecolopfical Enpfineerinpi 30: 78-88.

Hemond, H. F. 1980. Biogeochemistry of Thoreau’s Bog, Concord, Mass. Ecological
Monographs 50: 507-526.

Hey, D. L., and N. S. Philippi. 1995. Flood reduction through wetland restoration:
The Upper Mississippi River Basin as a case study. Restoration Ecology 3: 4-17.

Hunt, R. J., D. P. Krabbenhoft, and M. P. Anderson. 1996. Groundwater inflow mea-
surements in wedand systems. Water Resources Research 32: 495-507.

Ingram, H. A. P. 1983. Hydrology. In A. J. P. Gore, ed. Ecosystems of the World, Vol.
4A: Mires: Swamp, Bog, Fen, and Moor. Elsevier, Amsterdam, The Netherlands,
pp. 67-158.

Junk, W. J. 1982. Amazonian floodplains: Their ecology, present and potential use.
In D. O. Logofet and N. K. Luckyanov, eds.. Ecosystem Dynamics in Freshwater
Wetlands and Shallow Water Bodies, Vol. 1. SCOPE and UNEP Workshop, Center
of International Projects, Moscow, pp. 98-126.

Kantrud, H. A., J. B. Millar, and A. G. van der Valk. 1989. Vegetation of wetlands of
the prairie pothole region. In A. G. van der Valk, ed.. Northern Prairie Wetlands.
Iowa State University Press, Ames, pp. 132-187.

Keddy, P. A. 1992. Water level fluctuations and wetland conservation. In J. Kusler and
R. Smandon, eds.. Wetlands of the Great Lakes. Proceedings of an International
Symposium. Association of State Wetland Managers, Berne, NY, pp. 79-91.

Koreny, J. S., W. J. Mitsch, E. S. Bair, and X. Wu. 1999. Regional and local hydrology
of a constructed riparian wetland system. Wetlands 19: 182-193.

Korgen, B. J. 1995. Seiches. American Scientist July-August 1995: 330-341.

Leopold, L. B., M. G. Wolman, and J. E. Miller. 1964. Fluvial Processes in Geomor-
phology. W. H. Freeman, San Francisco. 522 pp.

Linacre, E. 1976. Swamps. In J. L. Monteith, ed.. Vegetation and the Atmosphere, vol.
2: Case Studies. Academic Press, London, pp. 329-347.

Lott, R. B., and R. J. Hunt. 2001. Estimating evapotranspiration in natural and con-
structed wetlands. Wetlands 2\-. 614-628.

Megonigal, J. P, W. H. Conner, S. Kroeger, and R. R. Sharitz. 1997. Aboveground
production in southeastern floodplain forests: A test of the subsidy-stress hypoth-
esis. Ecology 78: 370-384.

158 Chapter 4 Wetland Hydrology

Millar, J. B. 1971. Shoreline -are a as a factor in rate of water loss from small sloughs.
Journal of Hydrology 14: 259-284.

Mitsch, W. J. 1979. Interactions between a riparian swamp and a river in southern
Illinois. In R. R. Johnson and J. F. McCormick, tech, coords.. Strategies for the
Protection and Management of Floodplain Wetlands and Other Riparian Ecosys-
tems. Proceedings of the Symposium, Calaway Gardens, GA, December 11-13,
1978. General Technical Report WO-12, U.S. Forest Service, Washington, DG,
pp. 63-72.

Mitsch, W. J., and K. G. Ewel. 1979. Gomparative biomass and growth of cypress in
Florida wetlands. American Midland Naturalist 101: 417-426.

Mitsch, W. J., ed. 1989. Wetlands of Ohio’s Coastal Lake Erie: A Hierarchy of Systems.
NTIS, OHSU-BS-007, Ohio Sea Grant Program, Golumbus. 186 pp.

Mitsch, W. J., W. Rust, A. Behnke, and L. Lai. 1979. Environmental Observations of a
Riparian Ecosystem during Elood Season. Research Report 142, Illinois University
Water Resources Genter, Urbana. 64 pp.

Mitsch, W. J., and W. G. Rust. 1984. Tree growth responses to flooding in a bottom-
land forest in northeastern Illinois. E or est Science 30: 499-510.

Mitsch, W. J., and B. G. Reeder. 1992. Nutrient and hydrologic budgets of a Great
Lakes coastal freshwater wetland during a drought year. Wetlands Ecology and
Management 1(4): 211-223.

Mitsch, W. J., L. Zhang, G. J. Anderson, A. Altor, and M. Hernandez. 2005. Greating
riverine wedands: Ecological succession, nutrient retention, and pulsing effects.
Ecological Enpiineerinpt 25: 510-527.

Mitsch, W.J., A.M. Nahlik, P. Wolski, B. Bernal, L. Zhang, and L. Ramberg. 2010.
Tropical wedands: Seasonal hydrologic pulsing, carbon sequestration, and
methane emissions. Wetlands Ecology and Management 18: 573-586.

Mulholland, P. J., and E. J. Kuenzler. 1979. Organic carbon export from upland and
forested wedand watersheds. Limnolopiy and Oceanography 24: 960-966.

Nixon, S. W., and G. A. Oviatt. 1973. Ecology of a New England salt marsh. Ecological
Monographs 43: 463-498.

Odum, E. P. 1979. Ecological importance of the riparian zone. In R. R. Johnson and
J. F. McGormick, tech, coords.. Strategies for Protecdon and Management of
Floodplain Wetlands and Other Riparian Ecosystems. Proceedings of the Sympo-
sium, Gallaway Gardens, GA, December 11-13, 1978. General Technical Report
WO-12, U.S. Forest Service, Washington, DG, pp. 2-4.

Odum, H. T. 1971. Environment, Power and Society. John Wiley & Sons, New York.

Odum, W. E., E. P. Odum, and H. T. Odum. 1995. Nature’s pulsing paradigm. Estu-
aries 18: 547-555.

Penman, H. L. 1948. Natural evaporadon from open water, bare soil and grass. Pro-
ceedings of the Royal Society of London 93: 120-145.

Pride, R. W., F. W. Meyer, and R. N. Gherry. 1966. Hydrology of Green Swamp Area
in Central Florida. Report 42, Florida Division of Geology, Tallahassee. 137 pp.

Richardson, G. J. 1983. Pocosins: Vanishing wastelands or valuable wedands. Bio-
Science 33: 626-633.

References 159

Rosenberry, D. O., D. I. Stannard, T. C. Winter, and M. L. Martinez. 2004. Compar-
ison of 13 equations for determining evapotranspiration from a prairie wetland,
Cottonwood Lake Area, North Dakota, USA. Wetlands 24: 483M97.

Shjeflo, J. B. 1968. Evapotranspiration and the water budget of prairie potholes in
North Dakota. Professional Paper 585-B. U.S. Geological Survey, Washington,
DC. 49 pp.

Steever, E. Z., R. S. Warren, and W. A. Niering. 1976. Tidal energy subsidy and
standing crop production of Spartina alterniflora. Estuarine, Coastal and Marine
Science 4: 473-478.

Tutde, C. L., L. Zhang, and W. J. Mitsch. 2008. Aquatic metabolism as an indicator of
the ecological effects of hydrologic pulsing in flow-through wetlands. Ecolo£iical
Indicators 8: 795-806.

Twilley, R. R. 1982. Litter dynamics and organic carbon exchange in black man-
grove (Avicennia £erminans) basin forests in a Southwest Florida estuary. Ph.D.
dissertation. University of Florida, Gainesville.

U.S. Department of Interior (USDl), Bureau of Reclamation. 2001. Water Measure-
ment Manual, 3rd ed., revised reprint. U.S. Government Printing Office, Wash-
ington, DG, 327 pp. Available atwww.usbr.gov/pmts/hydraulics_lab/pubs/wmm/.

van der Valk, A. G., ed. 1989. Northern Prairie Wetlands. Iowa State University Press,
Ames. 400 pp.

Woo, M.-K., and T. G. Winter. 1993. The role of permafrost and seasonal frost in
the hydrology of nothern wetlands in North America. Journal ofHydrolojjy 141:
5-31.

Vernal Pools: A Community Profile. Biological Report 85 (7.11), U.S. Fish and Wildlife
Service, Washington, DG.

Chapter 5

Wetland Soils

Wetland soils, known as hydric soils, are defined as soils that formed under
conditions of saturation, flooding or pondin£i lonp; enough during the growing
season to develop anaerobic conditions in the upper part. They occur when
oxygen is cut off due to the presence of water, causing a predictable sequence of
chemically reduced conditions. Wetland soils can be organic soils or mineral
soils. Hydric mineral soils can be identified through hydric soil indicators such
as redox concentrations, redox depletions, and reduced soil matrices, often by
using a Munsell color chart. Redox potential is a useful measure of the degree to
which wetland soils are oxidized or reduced. Various chemical and biological
tranfiormations take place as coupled oxidation (e~ donor)-reduction
(e~ acceptor) reactions in wetland soils and occur in a predictable sequence
within predictable redox ranges.

lypes and Definitions

Wetland soils are both the medium in which many of the wedand chemical transfor-
mations take place and the primary storage of available chemicals tor most wetland
plants. They are often described as hydric soils, defined by the U.S. Department of
Agriculture’s Natural Resources Conservation Service (NRCS, 2010) as soils “that
formed under conditions of saturation, flooding or ponding long enough during the
growing season to develop anaerobic conditions in the upper part.” Wetland soils are
of two types: (1) mineral soils or (2) organic soils. All soils have some organic material,
but when a soil has less than 20 to 35 percent organic matter (on a dry-weight basis),
it is considered a mineral soil.

Organic soils and organic soil materials {peat, mucky peat, and muck) are defined
under either of two conditions of saturation:

1 . Soils are saturated with water for long periods or are artificially drained and,
excluding live roots, (a) have 18 percent or more organic carbon if the

161

162 Chapter 5 Wetland Soils

Figure 5.1 Percentage organic carbon required for a soil materiai to be called organic,
mucky modified minerai, or minerai soil material versus clay content. (Ftom NRCS, 2010)

mineral fraction is 60 percent or more clay, (b) have 12 percent or more
organic carbon if the mineral fraction has no clay, or (c) have a proportional
content of organic carbon between 12 and 18 percent if the clay content of
the mineral fraction is between 0 and 60 percent (Fig. 5.1); or

2. Soils are never saturated with water for more than a few days and have 20
percent or more organic carbon.

For an estimate of organic carbon when organic matter content is known,

%C„^ = %OAf/2 (5.1)

where

%C„r^ = percentage of organic carbon
% OM = percentage of organic matter

Any soil material that is not included in the preceding definition is considered
mineral soil material. Where mineral soils occur in wetlands, such as in some freshwater
marshes or riparian forests, they generally have a soil profile made up of horizons, or
layers. The upper layer of wetland mineral soils is often organic peat composed of
partially decayed plant materials.

Although the preceding definition of organic soil is applicable to many types of
wetlands, particularly to northern peatlands, peat^ a generic term for relatively unde-
composed organic soil material, is not usually that strictly defined. Most peats contain

Types and Definitions 163

less than 20 percent unburnable inorganic matter (and therefore usually contain more
than 80 percent burnable organic material, which is about 40 percent organic carbon).
Some soil scientists, however, allow up to 35 percent unburnable inorganic matter
(approximately 33 percent organic carbon), and commercial operations sometimes
allow 55 percent unburnable material (22 percent organic carbon). Muck is dehned
as sapric organic soil material with plant material so decomposed that identihcation of
plant forms is not possible. Its bulk density is generally greater than 0.2 g/cm^ and
more than peat.

Organic soils are different from mineral soils in four physicochemical features
other than the percentage of organic carbon (Table 5.1):

1 . Bulk density and porosity. Organic soils have lower bulk densities and higher
water-holding capacities than do mineral soils. Bulk density, defined as the
dry weight of soil material per unit volume, is generally 0.2 to 0.3 g/cm^
when the organic soil is well decomposed, although peatland soils composed
of Sphagnum moss can be extremely light, with bulk densities as low as
0.04 g/cm^. By contrast, mineral soil bulk density generally ranges between
1.0 and 2.0 g/cm^. Bulk density is low in organic soils because of their high
porosity, or percentage of pore spaces. Peat soils generally have at least 80
percent pore spaces and are thus 80 percent water by volume when flooded.
Mineral soils generally range from 45 to 55 percent total pore space,
regardless of the amount of clay or texture.

2. Hydraulic conductivity. Both mineral and organic soils have wide ranges of
possible hydraulic conductivities. Organic soils may hold more water than
mineral soils, but, given the same hydraulic conditions, they do not
necessarily allow water to pass through more rapidly. Hydraulic conductivity
can be predicted for some peatland soils from their bulk density or fiber
content, both of which can easily be measured (Fig. 5.2). In general, the
conductivity of organic peat decreases as the fiber content decreases through
the process of decomposition. Water can pass through fibric, or poorly

Table 5.1 Comparison of mineral and organic soils in wetlands

Minerai Soii

Organic Soil

Organic content (percent)

Less than 20 to 35

Greater than 20 to 35

Organic carbon (percent)

Less than 12 to 20

Greater than 12 to 20

Usuaiiy circumneutral

Acid

Buik density

High

Low

Porosity

Low (45-55%)

High (80%)

Hydrauiic conductivity

High (except for clays)

Low to high

Water hoiding capacity

Low

High

Nutrient avaiiabiiity

Generally high

Often low

Cation exchange capacity

Low, dominated by major cations

High, dominated by hydrogen ion

Typicai wetland

Riparian forest, some marshes

Northern peatland

164 Chapter 5 Wetland Soils

Bulk density, g/cm

Figure 5.2 Permeability of peatland sell as a function of fiber content and bulk density.
(After Verry and Beelter, 1979)

decomposed, peats 1,000 times faster than it can through more decomposed
sapric peats. The type of plant material that makes up the peat is also
important. Peat composed of the remains of grasses and sedges such as
Phra^mites and Carex, for example, is more permeable than the remains of
most mosses, including sphagnum. The hydraulic conductivity of peat can
vary over several orders of magnitude, showing a range almost as great as the
range for mineral soil between clay {k=Sx 10~^ cm/s) and sand
{k= 5 X 10“^ cm/s) (Table 5.2). There has been some disagreement over the
appropriate methods for measuring hydraulic conductivity in wetlands and
about whether Darcy’s law applies to flow through organic peat.

3. Nutrient availability. Organic soils generally have more minerals tied up in
organic forms unavailable to plants than do mineral soils. This follows from
the fact that a greater percentage of the soil material is organic. This does not
mean, however, that there are more total nutrients in organic soils; very
often, the opposite is true in wetland soils. For example, organic soils can be
extremely low in bioavailable phosphorus or iron content — enough to limit
plant productivity.

Organic Wetland Soil 165

Table 5.2 Typical hydraulic conductivity for wetland soils compared with other soii materials

Wetland or Soil Type

Hydraulic Conductivity,
k (cm/s X 10“®)

Reference

Northern Peatlands

Highly humified blanket bog, UK

0.02-0.006

Ingram (1967)

Fen, Russia

Slightly decomposed

500

Romanov (1968)

Moderately decomposed

Highly decomposed

Carex fen, Russia

0-50cm deep

310

Romanov (1968)

100-150cm deep

North American peatlands (general)

Fibric

>150

Verry and Boelter (1979)

Hemic

1.2-150

Sapric

<1.2

Coastal Salt Marsh

Great Sippewissett Marsh, Massachusetts

Hemond and Fifield (1982)

(vertical conductivity)

0-30cm deep

1.8

High permeability zone

2,600

Sand-peat transition zone

9.4

Nonpeat Wetland Soils

Cypress dome, Florida

Clay with minor sand

0.02-0.1

Smith (1975)

Sand

Okefenokee Swamp watershed, Georgia

3.4-834

Hyatt and Brook (1984)

Mineral Soils (general)

Clay

0.05

Limestone

5.0

Sand

5000

4. Cation exchan£;e capacity. Organic soils have a greater cation exchange

capacity, defined as the sum of exchangeable cations (positive ions) that a soil
can hold. Figure 5.3 summarizes the general relationship between organic
content and cation exchange capacity of soils. Mineral soils have a cation
exchange capacity that is dominated by the major metal cations (Ca^‘*‘, Mg^"'',
K"'', and Na"''). As organic content increases, both the percentage and the
amount of exchangeable hydrogen ions increase. For Sphagnum moss peat,
the high cation capacity may be caused by long-chain polymers of uronic acid
(Clymo, 1983).

Organic Wetland Soil

Organic soil is composed primarily of the remains of plants in various stages of
decomposition and accumulates in wetlands as a result of the anaerobic conditions

166 Chapter 5 Wetland Soils

Figure 5.3 Relationship between cation exchange capacity and organic content for wetland
soils. For low organic content (minerai soils), the cation exchange capacity is saturated by
metai cations; when organic content is high, the exchange capacity is dominated by hydrogen
ions. (After Gorham, 1967)

created by standing water or poorly drained conditions. Two of the more important
characteristics of organic soil, including soils commonly termed peat and muck, are
the botanical origin of the organic material and the degree to which it is decomposed.
Several of the properties that have been discussed, including bulk density, cation
exchange capacity, hydraulic conductivity, and porosity, are often dependent on these
characteristics. Therefore, it is often possible to predict the range of the physical
properties of an organic soil if the origin and state of decomposition can be observed
in the field or laboratory.

Botanical Origin

The botanical origin of the organic material can be (1) mosses, (2) herbaceous mate-
rial, and (3) wood and leaf litter. For most northern peadands, the moss is usually
Spha£inum, although several other moss species can dominate if the peatland is receiv-
ing inflows of mineral water. Organic soils can originate from herbaceous grasses such
as reed grass {Phrapjmites), wild rice {Zizania), and salt marsh cordgrass {Spartina),

Mineral Wetland Soil 167

or from sedges such as Carex and Cladium. Organic soils can also be produced in
freshwater marshes by plant fragments from several nongrass and nonsedge plants,
including cattails (Typha) and water lilies (Nymphaea). In forested wetlands, the peat
can be a result of woody detritus or leaf material or both. In northern peatlands, the
material can originate from birch (Betula), pine (Pinus), or tamarack (Larix), and in
southern deepwater swamps, the organic horizon can be composed of material from
cypress {Taxodium) or water tupelo (Nyssa) trees.

Decomposition

The state of decomposition, or humification, of wetland soils is the second key charac-
teristic of organic peat. As decomposition proceeds, albeit at a very slow rate in flooded
conditions, the original plant structure is changed physically and chemically until the
resulting material little resembles the parent material. As peat decomposes, bulk den-
sity increases, hydraulic conductivity decreases, and the quantity of larger (>1.5 mm)
fiber particles decreases as the material becomes increasingly fragmented. Chemically,
the amount of peat “wax,” or material soluble in nonpolar solvents, and lignin increase
with decomposition, whereas cellulose compounds and plant pigments decrease.
When some wetland plants, such as salt marsh grasses, die, the detritus rapidly loses
a large percentage of its organic compounds through leaching. These readily soluble
organic compounds are thought to be easily metabolized in adjacent aquatic systems.

Classification and Characteristics

Organic soils {histosols) are classified into four groups, the first three of which listed
here are considered hydric soils:

1. Saprists (muck). Two-thirds or more of the material is decomposed, and less
than one-third of plant fibers are identifiable.

2. Fibrists (peat). Less than one-third of material is decomposed, and more than
two-thirds of plant fibers are identifiable.

3. Hemists (mucky peat or peaty muck). Conditions fall between saprist and
fibrist soil.

4. Folists. Organic soils caused by excessive moisture (precipitation >
evapotranspiration) that accumulate in tropical and boreal mountains; these
soils are not classified as hydric soils because saturated conditions are the
exception rather than the rule.

Organic soil is generally dark in color, ranging from the dark black soils charac-
teristic of mucks such as those found in the Everglades in Florida to the dark brown
color of partially decomposed peat from northern bogs.

Mineral Wetland Soil

Mineral soils, when flooded for extended periods, develop certain characteristics that
allow for their identification. These characteristics are collectively called redoximorphic

168 Chapter 5 Wetland Soils

features^ defined as features formed by the reduction, translocation, and/or oxidation
of iron and manganese oxides (Vepraskas, 1995).

The development of redoximorphic features in mineral soils is mediated by micro-
biological processes. The rate at which they are formed depends on three conditions,
all of which must be present:

1 . Sustained anaerobic conditions

2. Sufficient soil temperature (5°C is often considered “biological zero,” below
which much biological activity ceases or slows considerably; see description of
biological zero and its importance to wetland science by Rabenhorst, 2005)

3. Organic matter, which serves as a substrate for microbial activity

Reduced Matrices and Redox Depletions

One characteristic of many hydric mineral soils that are semipermanently or perma-
nently flooded is the development of black, gray, or sometimes greenish or blue-gray
color as the result of a process known z.%gleying. This process, also known glcizMion^
is the result of the chemical reduction of iron (see “Iron and Manganese Transforma-
tions” in Chapter 6: “Wetland Biogeochemistry”). When soils are not saturated with
water, iron (ferric = Fe^"^) oxides are the principal chemicals that give the soil its typi-
cal red, brown, yellow, or orange color. Manganese (Mn^+ or Mn^+) oxides give the
soil a black color. When soils are flooded and become reduced, the iron is reduced to
a soluble form of iron (ferrous = Fe^'*’) and the manganese is reduced to its soluble
manganous (Mn^'*') form. These soluble forms of iron and manganese can be leached
out of the soil, leaving the natural (gray or black) color of the parent sand, silt, or
clay, called the matrix. A similar term used to describe these reduced soils is r^iox
depletions — iron is reduced and then depleted from the soil matrix. In a similar man-
ner, clay depletions occur when clay is selectively removed along root channels after
iron and manganese oxides have been depleted, only to redeposit as clay coatings on
soil particles below the clay depletions (Vepraskas, 1995).

Oxidized Rhizosphere

Another characteristic of some mineral wedand soils is the presence of an oxidized
rhizosphere (also called oxidized pore linings) that results from the capacity of
many hydrophytes to transport oxygen through aboveground stems and leaves to
below-ground roots (Fig. 5.4). Excess oxygen, beyond the root’s metabolic needs,
diffuses from the roots to the surrounding soil matrix, forming deposits of oxidized
iron along small roots. When a wetland soil is examined, these oxidized rhizosphere
deposits can often be seen as thin traces through an otherwise dark matrix.

Redox Concentrations

Mineral soils that are seasonally flooded, particularly by alternate wetting and dry-
ing, develop spots of highly oxidized materials called mottles or redox concentrations
(Fig. 5.5). Mottles and redox concentrations are orange/reddish-brown (because of

Mineral Wetland Soil 169

Oxidized

4hizosphere

(oxidized

pore

linings)

Figure 5.4 Formation of oxidized rhizospheres, or pore linings, around the roots of a wet-
iand plant caused by the transport of excess oxygen by wetland plants to their roots. When
the piant dies, pore linings of iron and manganese oxides often remain in the soil. (After
Vepraskas, 1995)

iron) or dark reddish-brown/black (because of manganese) spots seen throughout
an otherwise gray (gleyed) soil matrix and suggest intermittently exposed soils with
spots of iron and manganese oxides in an otherwise reduced environment. Mottles
are relatively insoluble, enabling them to remain in soil long after it has been drained.

Modern Nomenclature

A revised set of terms defining redoximorphic features has been devised by soil
scientists to describe indicators of hydric soils, or more properly, to identify an
aquic condition — the condition in which soils are saturated with water, are reduced,
and display redoximorphic features. The term aquic condition was introduced in
the early 1990s to better reconcile field techniques that used soil colors (e.g., iron

170 Chapter 5 Wetland Soils

concretion in matrix

soft Fe/Mn hard Fe/Mn

accumulations accumulations

nodules concretions

Figure 5.5 Different kinds ef redox concentrations, nr mottles, in soil peds (soii macropai^
tides), including nodules and concretions, iron masses in soil matrix (also called reddish
mottles), and pore finings on root channel (also called oxidized rhizospheres). (After
Vepraskas, 1995)

reduction or oxidation) with the former term aquic moisture regime — any soil that
was saturated with water and chemically reduced such that no dissolved oxygen was
present. The redoximorphic features that can be used to identify aquic conditions are
(Vepraskas, 1995):

1. Redox concentmtions. Accumulation of iron and manganese oxides (formerly
called mottles) in at least three different structures (Fig. 5.5):

a. Nodules and concretions. Firm to extremely firm irregularly shaped bodies
with diffuse boundaries

b. Masses. Formerly called reddish mottles

c. Pore linings. Formerly included oxidized rhizospheres (Figs. 5.4 and 5.5)

2. Redox depletions. Low-chroma (<2) bodies with high values (>4) including:

a. Iron depletions. Sometimes called gray mottles or gley mottles; these are
low-chroma bodies

b. Clay depletions. Contain less iron, manganese, and clay than adjacent soils

3. Reduced matrices. Low-chroma soils (because of presence of Fe^+) in situ
that change color if exposed to air and iron is oxidized to Fe^'*’

Mineral Hydric Soil Determination

In practice, the determination of whether a mineral soil is a hydric soil is a
complicated process, but it is often done by determining soil color relative to
a standard color chart called the Munsell® Soil Color Chart (Fig. 5.6a). Soils

Reduction/Oxidation in Wetland Soil 171

that contain low chromas (as indicated by the color chips on the left-hand side
of the color chart in Fig. 5.6b) indicate hydric soils. Soils that contain bright
reds, browns, yellows, or oranges are nonhydric. In general, a chroma of 2
or less on the Munsell color chart is necessary for a soil to be classified as
a hydric soil. These color charts are commonly used in the United States to
identify the presence of hydric soils for the delineation of wetlands.

(a) (b)

Figure 5.6 (a) Hydric soils can be identified by comparing the soil color with standard
soil color charts such as the Munsell Soil Color Chart shown here, (b) A representative
Munsell Soil Color Chart (lOYR in this case): The hue, given in the upper right-hand cor-
ner of the chart, indicates the relation to standard spectral colors, in this case yellow
(Y) and red (R). The value notation (vertical scale) indicates the soil lightness (darker
with lower value), and the chroma (horizontal scale) indicates the color strength or
purity, with grayer soils to the left. Chromas of 2 or less generally indicate hydric soils.

Reduction/Oxidation in Wetland Soil

All soils contain air and water in a mineral/organic matrix. When soils, whether min-
eral or organic, are inundated with water, anaerobic conditions usually result as water
fills the air spaces or soil pores. When water fills the pore spaces, the rate at which
oxygen can diffuse through the soil is drastically reduced. Diffusion of oxygen in an
aqueous solution has been estimated at 10,000 times slower than oxygen diffusion
through a porous medium such as drained soil. This low diffusion rate leads rela-
tively quickly to anaerobic, or reduced, conditions, with the time required for oxygen
depletion on the order of several hours to a few days after inundation begins (Fig. 5.7).
The rate at which the oxygen is depleted depends on the ambient temperature, the

172 Chapter 5 Wetland Soils

Figure 5.7 Sequence in time of transformations in soil after flooding, beginning with oxygen
depletion and followed by nitrate and then suifate reduction. Increases are seen in reduced
manganese (manganous), reduced iron (ferrous), hydrogen sulfide, and methane. Note the
gradual decrease in organic substrate (electron donor) and increases in availabie ammonium
(NH4''') and phosphate (P04^~) ions. The graph can aiso be interpreted as reiative concentra-
tions with depth in wetiand soils. (After Reddy and DeLaune, 2008)

availability of organic substrates for microbial respiration, and sometimes the chemical
oxygen demand from reductants such as ferrous iron. The resulting lack of oxygen
prevents plants from carrying out normal aerobic root respiration and strongly affects
the availability of plant nutrients and toxic materials in the soil. As a result, plants that
grow in anaerobic soils generally have some specific adaptations to this environment
(see Chapter 7).

It is not always true that oxygen is totally depleted from the soil water of wedands.
There is usually a thin layer of oxidized soil, sometimes only a few millimeters thick,
at the surface of the soil at the soil-water interface (Fig. 5.8). The thickness of this
oxidized layer is directiy related to four things:

1 . The rate of oxygen transport across the atmosphere-surface water interface

2. The small population of oxygen-consuming organisms present

3 . Photosynthetic oxygen production by algae within the water column

4. Surface mixing by convection currents and wind action

Even though the deeper layers of the wedand soils remain reduced, this thin oxi-
dized layer is often very important in the chemical transformadons and nutrient cycling
that occur in wedands. Oxidized ions such as Fe^"^, Mn'^"'', NOg^, and 804“ are found

Reduction/Oxidation in Wetland Soil 173

' Manganese, Iron, Sulfide, Redox potential,

1 ppm ppm cpm/g mv

Figure 5.8 Characteristics ef many wetland sails shewing a shailow oxidized soil layer over
a reduced soil layer. Also shown are soil profiles of reduced forms of manganese (sodium
acetate-extractable manganese), iron (ferrous iron), and sulfur (sulfide), and redox poten-
tial. (After Patrick and Delaune, 1972)

in this microlayer, whereas the lower anaerobic soils are dominated by reduced forms,
such as ferrous and manganous salts, ammonia, and sulfides. Because of the presence
of oxidized ferric iron (Fe^'*’) in the oxidized layer, the soil surface often is a brown
or brownish-red color, in contrast to the bluish-gray to greenish-gray color of the
reduced gleyed sediments, dominated by ferrous iron (Fe^'*‘).

Redox potentml, or oxidation-reduction potential, a measure of the electron pres-
sure (or availability) in a solution, is often used to further quantify the degree of elec-
trochemical reduction of wetland soils. Oxidation occurs not only during the uptake
of oxygen but also when hydrogen is removed (e.g., H2S ->■ + 2H~^) or, more gen-

erally, when a chemical gives up an electron (e.g., Fe^'*‘ ^Fe^"*" +e“). Reduction is the
opposite process of releasing oxygen, gaining hydrogen (hydrogenation), or gaining
an electron.

174 Chapter 5 Wetland Soils

Measuring Redox Potential

Redox potential can be measured in wetland soils and is a quantitative mea-
sure of the tendency of the soil to oxidize or reduce substances. When based
on a hydrogen scale, redox potential is referred to as and is related to the
concentrations of oxidants (ox) and reductants (red) in a redox reaction by the
Nernst equation:

Eh = E° + 2.3[RT/nF]log[ox]/{red] (5.2)

where

£° = potential of reference (mV)

R = gas constant = 81.987 cal deg“^ mol“^

T = temperature (°K)

n = number of moles of electrons transferred
F= Faraday constant = 23,061 cal/mole-volt

Redox potential can be measured with a platinum electrode (Fig. 5.9a, b),
which is easily constructed in the laboratory. Electric potential in units of mil-
livolts (mV) is measured relative to a hydrogen electrode (H+ -i- e — > H) or to
a calomel reference electrode. As long as free dissolved oxygen is present in
a solution, the redox potential varies little (in the range of -1-400 to -1-700 mV).
However, it becomes a sensitive measure of the degree of reduction

Figure 5.9 Design of (a, b) constructed redox and (c) possibie depioyment of muitiple
redox probes and reference eiectrode in a wetland for redox potential measurements,
((a), (b): After Faulkner et al., 1989)

12-gauge
'copper wire '

waterproof

epoxy

- 18-gauge _

platinum wire

silicon sealant
covered with
heatshrink tubing

glass tubing

. triple-distilled
mercury

Redox probes and reference |
electrode connected to
volt meter

Reduction/Oxidation in Wetland Soil 175

of wetland soils after oxygen disappears, ranging from +400 mV down to
-400mV. Flooding and/or redox conditions in pond and wetland soils can
be estimated by constructing platinum electrodes, using microplatinum
electrodes or steel rods. Normally a small tip of platinum (Pt) is connected
to a copper wire, which is then put in the soil, with the reference electrode
also put in the soil at a set distance (Fig. 5.9c). The potential between the
Pt tip and the reference electrode can be measured after the system has
stabilized, which takes sometimes up to two days.

As organic substrates in a waterlogged soil are oxidized (donate electrons), the
redox potential drops as a sequence of reductions (electron gains) takes place. Because
organic matter is one of the most reduced of substances, it can be oxidized when any
number of terminal electron acceptors is available, including O2, NOg^, Mn^"'', Fe^'*',
or 504“ . Rates of organic decomposition are most rapid in the presence of oxygen
and slower for electron acceptors such as nitrates and sulfates.

The oxidation of organic substrate is described by Equation 5.3, which illustrates
the organic substrate as an electron (e“) donor:

[CH20]w + WH2O ^ WCO2 + 4w£- + 4wH+ (5.3)

Various chemical and biological transformations take place as coupled oxidation
(e~ donor)-reduction (e“ acceptor) reactions. Equations 5.3 and 5.4 make one such
coupled reaction.

O2 + 4c- + 4H+ ^ 2H2O (5.4)

These transformations occur in a predictable sequence (Fig. 5.7), within pre-
dictable redox ranges to provide electron acceptors for this oxidation or decompo-
sition (Table 5.3).

The first and most common transformation is through aerobic oxidation when
oxygen itself is the terminal electron acceptor (Eq. 5.4) at a redox potential of between
400 and 600 mV.

Table 5.3 Oxidized and reduced forms of several elements and approximate
redox potentials for transformation

Element

Oxidized Form

Reduced Form

Approximate Redox
Potential for
Transformation (mV)

Nitrogen

(nitrate)

NjO, Nj, NH4+

250

Manganese

Mn'^’*' (manganic)

Mn^+ (manganous)

225

Iron

Fe^+ (ferric)

Fe^+ (ferrous)

-1-100 to -100

Sulfur

304= (sulfate)

S= (sulfide)

-100 to -200

Carbon

CO2 (carbon dioxide)

CFI4 (methane)

Below -200

176 Chapter 5 Wetland Soils

One of the first reactions that occur in wetland soils after they become anaerobic
(i.e., the dissolved oxygen is depleted) is the reduction of NOg” (nitrate) first to N02^
(nitrite) and ultimately to N2O (nitrous oxide) or N2 (nitrogen gas); nitrate becomes
an electron acceptor at a redox potential of approximately 250 mV:

2NO3 + IQe- + 12H+ ^ N2 + 6H2O (5.5)

As the redox potential continues to decrease, manganese is transformed from man-
ganic to manganous compounds at about 225 mV:

Mn02 + 2<j- + 4H+ ^ Mn^+ + 2H2O (5.6)

Iron is transformed from ferric to ferrous form at about +100 to -100, while
sulfates are reduced to sulfides at -100 to -200 mV:

Fe(OH)3 + + 3H+ ^ p2+ + 3H2O (5.7)

SO= + 8e-+9H+ ^HS-+4H20 (5.8)

Finally, under the most reduced conditions, the organic matter itself (or car-
bon dioxide) becomes the terminal electron acceptor below -200 mV, producing
low-molecular-weight organic compounds and methane gas, as, for example,

C02 + 8e- + 8H+ ^CH4 + 2H20 (5.9)

These redox potentials are not precise thresholds, because pH and temperature
are also important factors in the rates of transformation. These major chemical trans-
formations and others related to the nitrogen, sulfur, and carbon cycles are discussed
in the next chapter.

Recommended Readings

Basldn, Y. 2005. Under Ground: How Creatures of Mud and Dirt Shape Our World.
Washington, DC: Island Press.

Richardson, J. L., and M. J. Vepraskas. 2001. Wetland Soils: Genesis, Hydrology, Land-
scapes, and Classification. Boca Raton, FT: CRC Press.

Natural Resources Conservation Service (NRCS). 2010. Field Indicators of Hydric
Soils in the United States, Version 7.0. L. M. Vasilas, G. W. Gurt, G. V. Noble,
eds. USDA, NRGS in cooperation with the National Technical Gommittee for
Hydric Soils, 45 pp.

References

Glymo, R. S. 1983. Peat. In A. J. P. Gore, ed.. Ecosystems of the World, Vol. 4A: Mires:
Swamp, Bo£!, Fen, and Moor. Elsevier, Amsterdam, The Netherlands, pp. 1 59-224.

References 177

Faulkner, S. P., W. H. Patrick, Jr., and R. P Gambrell. 1989. Field techniques for
measuring wetland soil parameters. Soil Science Society of America Journal 53:
883-890.

Gorham, E. 1967. Some Chemical Aspects of Wetland Ecology. Technical Memorandum
90, Gommittee on Geotechnical Research, National Research Gouncil of Ganada,
pp. 2-38.

Hemond, H. F., and J. L. Fifield. 1982. Subsurface flow in salt marsh peat: A model
and field study. Limnolopiy and Oceanojjraphy 27: 126-136.

Hyatt, R. A., and G. A. Brook. 1984. Groundwater flow in the Okefenokee Swamp
and hydrologic and nutrient budgets for the period August, 1981 through July,
1982. In A. D. Gohen, D. J. Gasagrande, M. J. Andrejko, and G. R. Best, eds..
The Okefenokee Swamp: Its Natural History, Geology, and Geochemistry. Wetland
Surveys, Los Alamos, NM, pp. 229-245.

Ingram, H. A. P. 1967. Problems of hydrology and plant distribution in mires. Journal
of Ecology 55: 711-724.

Natural Resources Gonservation Service (NRGS). 2010. Field Indicators of Hydric
Soils in the United States, Version 7.0. L. M. Vasilas, G. W. Gurt, G. V. Noble,
eds. USDA, NRGS in cooperation with the National Technical Gommittee for
Hydric Soils, 45 pp.

Patrick, W. H., Jr., and R. D. Delaune. 1972. Gharacterization of the oxidized and
reduced zones in flooded soil. Proceedinpis of the Soil Science Society of America 36:
573-576.

Rabenhorst, M. G. 2005. Biological zero: A soil temperature concept. Wetlands 25:
616-621.

Reddy, K. R. and R. D. DeLaune. 2008. Bio^eochemistry of Wetlands. GRG Press,
Boca Raton, FT, 774 pp.

Romanov, V. V. 1968. Hydrophysics of Bopis. Translated from Russian by N. Kaner;
edited by Prof. Heimann. Israel Program for Scientific Translation, Jerusalem.
Available from Glearinghouse for Federal Scientific and Technical Information,
Springfield, VA. 299 pp.

Smith, R. G. 1975 . Hydrogeology of the experimental cypress swamps. In H. T. Odum
and K. G. Ewel, eds.. Cypress Wetlands for Water Manap/ement, Recycling and
Conservation. Second Annual Report to NSF and Rockefeller Foundation, Genter
for Wetlands, University of Florida, Gainesville, pp. 114-138.

Vepraskas, M. J. 1995. Redoximorphic Features for Identifyinp; Aquic Conditions. Tech-
nical Bulletin 301, North Garolina Agricultural Research Service, North Garolina
State University, Raleigh. 33 pp.

Verry, E. S., and D. H. Boelter. 1979. Peatland hydrology. In P. E. Greeson, J. R.
Glark, and J. E. Glark, eds.. Wetland Functions and Values: The State of Our
Understanding . American Water Resources Association, Minneapolis, MN,
pp. 389-402.

Chapter S

Wetland Biogeochemistry

Wetland biqgeochemistry features a combination of many chemical
transformations and chemical transport processes. Many transformations of
nitropfen, sulfur, iron, manpfanese, carbon, phosphorus and other chemicals
occur in wetlands as a result of the combination of both aerobic and anaerobic
conditions in proximity. Wetlands can be sources, sinks, or transformers of
nutrients but are valued especially for their ability to be sinks for nutrients.

Some tranformations cause toxic conditions, as with the production of
hydropten sulfide, whereas others, such as sedimentation, denitrification, and
carbon sequestration, improve water quality and improve the planefs carbon
balance. Still other processes from wetlands allow emissions of greenhouse ^ases
to the atmosphere. Many transformations in wetlands, especially in the
nitropfen, sulfur, and carbon cycles, are mediated by microbial populations that
are adapted to the anaerobic environment, while many other processes, such as
those in the phosphorus cycle, are chemical and physical. Wetlands are often
coupled to adjacent ecosystems such as by exporting vital orpfanic carbon to
downstream aquatic ecosystems.

The transport and transformation of chemicals in ecosystems, known as bio^eochemical
cycling, involve a great number of interrelated physical, chemical, and biological pro-
cesses. The diverse hydrologic conditions in wetlands discussed in Chapter 4 and the
soil types described in Chapter 5 both markedly influence biogeochemical processes.
These processes result not only in changes in the chemical forms of materials but also in
the spatial movement of materials within wetlands, as in water-sediment exchange and
plant uptake, and with surrounding ecosystems, as in organic exports. These processes,
in turn, determine overall wetland productivity. The interrelationships among hydrol-
ogy, the physiochemical environment, and wetland biota were already summarized in
Figure 4.1.

179

180 Chapter 6 Wetland Biogeochemistry

Figure 6.1 Components of a wetland nutrient budget, including inflows, outflows, and
intrasystem cycling.

The biogeochemistry of wetlands can be divided into (1) intrasystem cycling
through various transformation processes and (2 ) the exchange of chemicals between
a wetland and surrounding waters, landscapes, and atmosphere (Figs. 6.1 and 6.2).
Although no transformation processes are unique to wetlands, the permanent to inter-
mittent flooding of these ecosystems causes certain processes to be more dominant in
wetlands than in either upland or deep aquatic ecosystems. For example, while anaer-
obic, or oxygen-less, conditions are sometimes found in other ecosystems, they prevail
in wetlands. Wedand soils are characterized by waterlogged conditions during part
or all of the year, which produce reduced conditions, which, in turn, have a marked
influence on several biochemical transformations unique to anaerobic conditions.

This intrasystem cycling, along with hydrologic conditions, influences the degree
to which chemicals are transported to or from wetlands. An ecosystem is consid-
ered biogeochemically open when there is an abundant exchange of materials with
its surroundings. When there is little movement of materials across the ecosystem
boundary, it is biogeochemically closed. Wetlands can fall into either category. For
example, wetlands such as bottomland forests and tidal salt marshes have a signifi-
cant exchange of minerals with their surroundings through river flooding and tidal
exchange, respectively. Other wetlands such as ombrotrophic bogs and cypress domes

Figure 6.2 A wetland can serve as an (a) inerganic nutrient sink, (b) source of total nutri-
ents, and (c) transformer of inorganic nutrients to organic nutrients.

181

182 Chapter 6 Wetland Biogeochemistry

have little material exchange except for precipitation and gases that pass into or out
of the ecosystem. These latter systems depend more on intrasystem cycling than on
throughput for their chemical supplies.

Wetlands serve as sources, sinks, or transformers of chemicals or nutrients, depend-
ing on the wetland type, the hydrologic conditions, and the length of time the wetland
has been subjected to chemical loadings (Fig. 6.2). When wetlands serve as sinks for
certain chemicals (Fig. 6.2a), the long-term sustainability of that situation depends on
the hydrologic and geomorphic conditions, the spatial and temporal distribution of
chemicals in the wetland, and the ecosystem succession. Wetlands can become satu-
rated in certain chemicals after a number of years, particularly if inflows are high, and
become sources (Fig. 6.2b) or transformers (Fig. 6.2c) of chemicals.

The Nitrogen Cycle

The nitrogen cycle (Fig. 6.3) is one of the most important and studied chemical
cycles in wetlands. Nitrogen appears in a number of oxidation states in wetlands,
several of which are important in a wetland’s biogeochemistry. Nitrogen is often
the most limiting nutrient in flooded soils, whether the flooded soils are in natural
wetlands or on agricultural wetlands, such as rice paddies. Nitrogen is considered
one of the major limiting factors in coastal waters, making the nitrogen dynamics
in coastal wetlands particularly significant, although this universal belief in nitrogen
limitation in coastal wetlands has been challenged (e.g., Sundareshwar et al., 2005).
Because of the presence of anoxic conditions in wetlands, microbial denitrification of
nitrates to gaseous forms of nitrogen in wetlands and their subsequent release to the
atmosphere remain one of the more significant ways in which nitrogen is lost from
the lithosphere and hydrosphere to the atmosphere. Nitrates serve as one of the first
terminal electron acceptors in wetland soils in this situation after the disappearance
of oxygen (see Table 5.3), making them an important chemical in the oxidation of
organic matter in wetlands.

Nitrogen transformations in wetlands (Fig. 6.3) involve several microbiologi-
cal processes, some of which make the nutrient less available for plant uptake. The
ammonium ion (NH4+), with a nitrogen oxidation state of -3, is the primary form
of mineralized nitrogen in most flooded wetland soils, although much nitrogen can
be tied up in organic forms in highly organic soils. The presence of an oxidized zone
over the anaerobic or reduced zone is critical for several of the pathways.

Nitrogen Mineralization

Nitrogen minemlization refers to a series of biological transformations that converts
organically bound nitrogen to ammonium nitrogen as the organic matter is being
decomposed and degraded. This pathway occurs under both anaerobic and aerobic
conditions and is often referred to as ammonification. Typical formulas for the min-
eralization of a simple soluble organic nitrogen (SON) compound, urea, are given as

NH2CONH2 + H2O ^ 2NH3 -I- CO2
NH3 -I- H2O ^ NH4 -I- OH“

(6.1)

(6.2)

The Nitrogen Cycle 183

KEY

dinitrogen

N20

nitrous oxide

NH3

ammonia

nhJ

ammonium ion

NO 2
NO3

nitrite

nitrate

SON

soluble organic N

NH3

Volatilization

Nj N2O

Organic N — •

-I- -t-

OrganicN-*. 30n_^nh;-

Nitrification

runoff,

leaching

Organic N ■

4 4 diffusion

Upward diffusion *

I I Mitrate reduction ^

SON-^NHJ:^^— ^

Denitrification

Plant uptake

Figure 6.3 The nitrogen cycle in wetlands. Major pathways illustrated are nitrogen fixa-
tion, ammonia volatiiization, nitrification, denitrification, piant uptake, dissimilatery nitrate
reduction to ammonia (DNRA), and anammox (anaerobic ammonium oxidation)

Ammonia Transformations and Nitrification

Once the ammonium ion (NH4"'') is formed, it can take several possible pathways. It
can be absorbed by plants through their root systems or by anaerobic microorganisms
and converted back to organic matter. Under high-pH conditions (pH >8 ), a common
occurrence in marsh waters with excessive algal blooms, the ammonium ion can be
converted to NH3, which is then released to the atmosphere through volatilization.
The ammonium ion can also be immobilized through ion exchange onto negatively
charged soil particles. Because of the anaerobic conditions in wetland soils, ammonium
would normally be restricted from further oxidation and would build up to excessive
levels were it not for the thin oxidized layer at the surface of many wetland soils.
The gradient between high concentrations of ammonium in the reduced soils and low
concentrations in the oxidized layer causes an upward diffusion of ammonium, albeit
very slowly, to the oxidized layer. In this aerobic environment, ammonium nitrogen
can be oxidized through the process of nitrification in two steps by Nitr osomonas sp.:

2NH| + 3O2 ^ 2NO2 + 2H2O + 4H+ + energy

(6.3)

184 Chapter 6 Wetland Biogeochemistry

and by Nitrobacter sp.:

2NO2 + O2 ^ + energy (6.4)

Nitrification can also occur in the oxidized rhizosphere of plants, where adequate
oxygen is often available to convert the ammonium nitrogen to nitrate nitrogen.

Nitrate Transformations and Denitrification

Nitrate (NO3 ), as a negative ion rather than the positive ammonium ion, is not subject
to immobilization by negatively charged soil particles and is thus much more mobile in
solution. Ifit is not assimilated immediately by plants or microbes {assimilatory nitrate
reduction) or is lost through groundwater flow stemming from its rapid mobility, it has
the potential to undergo dissimilatory nitrogenous oxide reduction^ a term that refers
to several pathways of nitrate reduction. It is called dissimilatory because the nitrogen
is not assimilated into a biological cell. The most prevalent are reduction to ammonia
and denitrification.

Denitrification, carried out by facultative bacteria under anaerobic conditions,
with nitrate acting as a terminal electron acceptor, results in the loss of nitrogen as
it is converted to gaseous molecular nitrogen (N2) with some small fraction to nitrous
oxide (N2O):

C6H12O6 + 4N03~ ^ 6CO2 + 6H2O + 2N2 (6.5)

Denitrification is a significant path of nitrogen loss from most kinds of wetlands,
including salt marshes, freshwater marshes, forested wetlands, and rice paddies. Den-
itrification is inhibited in acid soils and peat and is therefore thought to be of less
consequence in northern peatlands. As illustrated in Figure 6.3, the entire process
occurs after (1) ammonium nitrogen diffuses to the aerobic soil layer, (2) nitrification
occurs, (3) nitrate nitrogen diffuses back to the anaerobic layer, and (4) denitrifica-
tion, as described in Equation 6.5, occurs. The diffusion rates of the ammonium ion
to the aerobic soil layer and the nitrate ion to the anaerobic layer are governed by the
concentration gradients of the ions. There is generally a steep gradient of ammonium
between the anaerobic and aerobic layers. Nevertheless, because nitrate diffusion rates
in wetland soils are seven times faster than ammonium diffusion rates, ammonium dif-
fusion and subsequent nitrification appear to limit the entire process of nitrogen loss
by denitrification.

Given an adequate supply of nitrate-nitrogen, the next most significant factor that
affects denitrification appears to be temperature. This pattern is clearly shown in a sum-
mary of denitrification measurements in created riverine wetlands in Ohio from 2004
through 2009 (Fig. 6.4). The highest nitrate-nitrogen concentrations in the wedands
are January through June (winter and spring), and the highest water temperatures
are in July through September (summer). Denitrification peaks in June, then, after
lower numbers in the early summer (due to low concentrations of nitrate nitrogen),
it peaks again in September, when temperatures remain high despite continued low
levels of nitrate-nitrogen. The strongest regression for denitrification for the multiyear
comparison was with water temperature (Fig. 6.5).

Figure 6.4 Seasonal patterns of denitrification and associated environmental variables for
two 1-ha created riverine wetlands in Ohio, summarized for studies in 2004, 2005, 2008,
and 2009. (a) Monthly averaged nitrate-nitrogen concentrations and water temperature of
inflowing water to the wetlands; (b) average ± standard error of monthly denitrification mea-
surements for open water, deepwater marsh (DM), shallow marsh (SM), and edge of wetland.
(Ftom Song et al., 2014; reprinted with permission, Elsevier.)

185

Water Temp (°C)

186 Chapter 6 Wetland Biogeochemistry

Figure 6.5 Relationship between denitrification and water temperature for the wetland
described in Figure 6.4. (Ftom Song et al. 2014); reprinted with permission, Eisevier

There are two gaseous products of denitrification — dinitrogen (N2) and nitrous
oxide (N2O). The predominant gas that usually results from denitrification in most
wetlands is N2, and that is no environmental issue with an atmosphere already having
80 percent N2. However, nitrous oxide is one of the so-called greenhouse gases that
could cause climate change, so any attempt to design wetlands for nitrate removal
should recognize this and understand conditions that minimize nitrous oxide produc-
tion in favor of dinitrogen production. Hernandez and Mitsch (2006, 2007) found
lower nitrous oxide fluxes in the spring in pulse -flooded conditions in the higher
marshes compared to steady flow conditions in the same Ohio riverine wetlands the
next year (Fig. 6.6). The rate is probably limited by the lack of nitrates in the per-
manently flooded soils. Nitrous oxide production was highest when soil tempera-
tures were greater than 20°C in the summer months (Fig. 6.6). In addition, wet-
land plants appeared to increase nitrous oxide emissions when sites were flooded but
not when soils were exposed. Overall, the amount of nitrogen emitted as nitrous
oxide in these riverine wetlands as a percentage of the total nitrogen released via
denitrification was quite small. This study suggests that nitrous oxide emissions and
nitrous oxide/dinitrogen gas ratios (N2O/N2) in denitrification are higher on the
aerobic/anaerobic edges of wetlands than in the more anaerobic middle. It is rea-
sonable to conclude then that if nitrate-nitrogen is denitrified in more aerobic farm
fields, ditches, streams and rivers, and even downstream coastal waters rather than in
wetlands, higher nitrous oxide emissions would result from those systems than from
the wetlands (Hernandez and Mitsch, 2006, 2007). Thus wetlands may not be the

The Nitrogen Cycle 187

Figure 6.6 Seasonal nitrous oxide fluxes under different hydrologic conditions and along
(a) dry edge, (b) high marsh (saturated soils with some standing water), and (c) low marsh
(continuous standing water) in the freshwater experimental marshes in central Ohio.
Numbers indicate number of flux measurements. (From Hernandez and Mitsch, 2006)

cause of additional nitrous oxide emissions; created and restored wetlands actually may
decrease the overall nitrous oxide emissions on a landscape scale.

Nitrogen Fixation

Nitrogen fixation results in the conversion of N2 gas to organic nitrogen through the
activity of certain organisms in the presence of the enzyme nitrogenase. It may be
the source of significant nitrogen for some wetlands. Nitrogen fixation, which is car-
ried out by certain aerobic and anaerobic bacteria and blue-green algae, is favored
in low oxygen conditions because nitrogenase activity is inhibited by high oxygen.

188 Chapter 6 Wetland Biogeochemistry

In wetlands, nitrogen fixation can occur in overlying waters, in the aerobic soil layer,
in the anaerobic soil layer, in the oxidized rhizosphere of the plants, and on the leaf and
stem surfaces of plants. Bacterial nitrogen fixation can be carried out by nonsymbiotic
bacteria, by symbiotic bacteria of the genus Rhizobium, or by certain actinomycetes.
Bacterial fixation is the most significant pathway for nitrogen fixation in salt marsh
soils, while nitrogen-fixing bacteria are virtually absent from the low-pH peat of north-
ern bogs. Cyanobacteria (blue-green algae) are common nitrogen fixers in wedands,
occurring in flooded delta soils in Louisiana, in northern bogs, and in rice cultures.

Dissimilatory Nitrate Reduction to Ammonia

Because conversion of nitrate-nitrogen to dinitrogen and nitrous oxide is consid-
ered to be the primary transformation of nitrates in anaerobic soils, an additional
process whereby nitrate -nitrogen is transformed in anaerobic conditions is often over-
looked (Megonigal et ah, 2004). The process — called dissimilatory nitrate reduction
to ammonia (DNRA) — occurs as follows, with mobile nitrates as the initial form of
nitrogen and less-mobile ammonium as the product.

NOs" + 4H2 + 2H+ ^ 3H2O + NH4+ (6.6)

The process yields energy to the many microorganisms capable of carrying out
this process. The bacteria can be anaerobic, aerobic, or facultative. In some cases,
nitrate reduction can be a more significant pathway than the other dissimilatory
nitrate loss — denitrification. Studies have supported the concept that high availability
of organic carbon and/or low nitrate concentrations favors DNRA over denitrification
(Megonigal et ah, 2004).

Anammox

Anammox (for anaerobic ammonium oxidation) involves nitrite-nitrogen (rather than
nitrate-nitrogen as originally thought) as the oxidant:

NO- +NH+ ^ 2H2O + N2 (6.7)

Few studies have definitively determined the importance of anammox in the
cycling of nitrogen in natural or created wetlands, but it does appear that this process
may be more important in wedands where denitrification is limited by lack of organic
carbon (Megonigal et ah, 2004). Erler et al. (2008) found anammox contributed
up to 24 percent of the dinitrogen production in a surface flow treatment wetland.
Ligi et al. (2015) detected bacterial genes that are specific to organisms capable
of anammox from soils samples taken from the Ohio created wetlands described
above. They suggested that anammox converting ammonium to dinitrogen gas may
compensate for the relatively low rates of denitrification reported at these wetlands
(Mitsch et al., 2012; Song et al., 2014).

The Nitrogen Cycle 189

The Nitrogen Cycle, Wetlands, and Hypoxia

Humans have essentially doubled the amount of nitrogen entering the
land-based nitrogen cycle through fertilizer manufacturing, increased use of
nitrogen-fixing crops, and fossil fuel burning (Galloway et al., 2003; Doering
et al., 2011). Significant amounts of this excess nitrogen are transported
as nitrate-nitrogen to rivers and streams, leading to eutrophication and
episodic and persistent hypoxia (dissolved oxygen <2 mg/L) in coastal waters
worldwide. For example, a hypoxic zone that currently averages close to
14,350 km^ reappears annually in the Gulf of Mexico (Figs. 6.7 and 6.8),
caused almost certainly by excessive nitrogen coming from farm fields in
from the Mississippi-Ohio-Missouri (MOM) river basin 1,000 km to the north
of the gulf. The extent of the hypoxia was much smaller than that area in the
late 1980s. The federal government decreed in 2000 and then again in 2008
that the hypoxia should be no larger than 5000 km^ (Mississippi River/Gulf
of Mexico Watershed Nutrient Task Force, 2008).

Many options were investigated for controlling nutrient flow into the gulf
by research teams in the late 1990s (e.g., Mitsch et al., 2001). In the end,
there were the general approaches that involve either revision of agronomic
approaches or wetland creation and riparian restoration that make the most
sense (Fig. 6.9). Two million ha of restored and created wetlands and restored
riparian buffers were recommended as necessary to provide enough denitrifi-
cation to substantially reduce the nitrogen entering the Gulf of Mexico (Mitsch
et al., 2001, 2005; Mitsch and Day, 2006). The anaerobic process of deni-
trification in wetlands was a particularly important process recognized in this
recommendation. Two million hectares of wetlands is less than 1 percent of
the Mississippi River Basin. Interestingly, Hey and Phillipi (1995) found that
a similar scale of wetland restoration would be required in the Upper Missis-
sippi River Basin to mitigate the effects of very large and costly floods, such
as the one that occurred in the summer of 1993 in the Upper Mississippi
River Basin.

Murphy et al. (2013) looked at the trend of nitrate-nitrogen in the Missis-
sippi River for the 30-year period of 1980 to 2010 and found that, although
two states (Iowa and Illinois) with the highest nitrate-nitrogen in the 1980s
had 11 to 15 percent reductions in nitrate-nitrogen concentrations and load-
ing over the 30 years, other sites on the river had 8 to 55 percent increases
in nitrate loading over that same period, essentially overshadowing the mod-
est decreases from Iowa and Illinois. The flux of nitrate-nitrogen to the Gulf
of Mexico increased by 14.5 percent over those 30 years and was at an
all-time high in 2010; the concentrations of nitrate-nitrogen entering the gulf
increased by 19 percent over those years. In the early 2000s, there was

a. 2011

-95 -94 -93 -92 -91 -90 -89

2012

-95 -94 -93 -92 -91 -90 -89

Bottom dtssotved
oxygen <2nr>g/L

Figure 6.7 Extent of hypoxic conditions in Guif of Mexico in summers of (a) 2011, (b)
2012, and (c) 2013. Shaded area indicates where guif waters are iess than 2mg/L in
dissolved oxygen. The hypoxia covered 17,520, 7,500, and 15,000 km^ in those three
years, respectively. The smaller hypoxia area in 2012 may have been due to an exten-
sive drought in the Midwestern USA that year that led to reduced Mississippi River
flows. Source: N. Rabalais, Louisiana Universities Marine Consortium, and NOAA, Cen-
ter for Sponsored Coastal Ocean Research.

190

25,000

Figure 6.8 The extent of mid-summer Guif of Mexico hypoxia from 1985 through 2014.
The average size over the iast five years (2010-2014) is shown to be about 14,350 km^.
Also shown is the action plan goal of 5,000 km^ set by a government task force in 2000
and reaffirmed in its action plan of 2008 (Mississippi River/Gulf of Mexico Watershed
Nutrient Task Force, 2008) Source: N. Rabalais, Louisiana Universities Marine Consor-
tium, and NOAA, Center for Sponsored Coastal Ocean Research.

fanning technigues

restored

bottomland

forest

created
wetland
intercepting
tile drainage

Figure 6.9 Sketch of strategy for wetland restoration and better farming practices in
an agricultural setting to improve water quality in the midwestern United States, par-
ticularly to control nitrate-nitrogen to protect the downstream Gulf of Mexico. Mitsch
et al. (2001, 2005) recommended 2 million ha of created and restored wetlands and
riparian forest buffers in the Mississippi River Basin to intercept surface and subsui^
face drainage coming from agricultural nonpoint sources. (Ftom Mitsch et al., 2001,
2005)

191

192 Chapter 6 Wetland Biogeochemistry

great anticipation in the federal government that the loading rates were going
to diminish due to the implementation of best management practices, includ-
ing created and restored wetlands. It simply has not happened. David et al.
(2013) argue that the causes are both biophysical and social within the agri-
cultural industry of the midwestern United States.

Iron and Manganese Transformations

Below the reduction of nitrate on the redox potential scale comes the reduction of
manganese and iron (see Equations 5.6 and 5.7 in Chapter 5). Iron and manganese
are among the most abundant minerals on Earth, and are found in wedands primarily
in their reduced forms (ferrous and manganous, respectively; see Table 5.3). Both are
more soluble and more readily available to organisms in those forms. Manganese is
reduced slightly before iron on the redox scale, but otherwise it behaves similarly to
iron. The direct involvement of bacteria in the reduction of manganic oxide (Mn02)
has been questioned by some researchers, although several experiments have shown
the generation of energy by the bacterial reduction of oxidized manganese (Laan-
broek, 1990).

Iron can be oxidized from reduced ferrous iron to the insoluble ferric form by
chemosynthetic bacteria in the presence of oxygen:

4Fe2+ -I- 02iaq) + 4H+ ^ 4Fe^+ -I- 2H2O (6.8)

Although this reaction can occur nonbiologically at neutral or alkaline pH, micro-
bial activity has been shown to accelerate ferrous iron oxidation by a factor of 10^ in
coal mine drainage water (Singer and Stumm, 1970). A similar type of bacterial process
is believed to exist for manganese .

Iron bacteria are thought to be responsible for the oxidation to insoluble fer-
ric compounds of soluble ferrous iron that originated in anaerobic groundwaters in
northern peatland areas. These “bog-iron” deposits form the basis of the ore that has
been used in the iron and steel industry. Iron in its reduced ferrous form causes a
gray-green coloration (gleying) of mineral soils instead of the normal red or brown
color in oxidized conditions caused by ferric hydroxide [Fe(OH)3]. This appearance
gives a relatively easy field check on the oxidized and reduced layers in a mineral soil
profile.

Iron and manganese in their reduced forms can reach toxic concentrations in
wetland soils. Ferrous iron, diffusing to the surface of the roots of wetland plants, can
be oxidized by oxygen leaking from root cells, immobilizing phosphorus and coating
roots with an iron oxide, and causing a barrier to nutrient uptake.

The Sulfur Cycle 193

The Sulfur Cycle

Sulfur, as the fourteenth most abundant element in the Earth’s surface, occurs in sev-
eral different states of oxidation in wetlands. Like nitrogen, it is transformed through
several pathways that are mediated by microorganisms (Fig. 6.10). Sulfur is rarely
present in such low concentrations that it is limiting to plant or animal growth in wet-
lands. The release of the reduced form of sulfur, sulfide (S ^), when wetland sediments
are disturbed causes the odor familiar to those who carry out research in wetlands — the
smell of rotten eggs as hydrogen sulfide (H2S). On the redox scale, sulfur com-
pounds are the next major electron acceptors after nitrates, iron, and manganese, with

KEY

H2S

hydrogen sulfide

S -2

elemental sulfur

SO4

sulfate

SO2

sulfur dioxide

FeS^,

DMS

FeS Iron sulfide (pyrite)

dimethyl sulfide

Figure 6.10 The sulfur cycle in wetlands. Majnr pathways illustrated are sulfur oxidation,
sulfate reduction, iron sulfide production, sulfate absorption and leaching, and hydrogen
sulfide emissions.

194 Chapter 6 Wetland Biogeochemistry

reduction occurring at about -100 to -200mV on the redox scale (see Table 5.3). The
most common oxidation states (valences) for sulfur in wetlands are:

Form

Valence

(sulfide)

-2

S (elemental sulfur)

S2O3 (thiosulfate)

S04^ (sulfate)

Sulfate Reduction

Sulfate reduction can take place as assimilatory sulfate reduction in which certain
sulfur-reducing obligate anaerobes, such as Desulfovibrio bacteria, utilize the sulfates
as terminal electron acceptors in anaerobic respiration:

4H2 + SO= ^ H2S + 2H2O + 2OH- (6.9)

This sulfate reduction can occur over a wide range of pH, with the highest rates
prevalent near neutral pH.

There have been a few measurements of the rate at which hydrogen sulfide is
produced in and released from wetlands, and those measurements have ranged over
several orders of magnitude. It can be safely generalized that saltwater wetlands have
higher rates of sulfide emission per unit area than do freshwater weriands, where sulfate
ions are much less abundant (~2700mg/L in sea water; ~10mg/L in fresh water).
Sulfur can also be released to the atmosphere as organic sulfur compounds, especially
as dimethyl sulfide (DMS), (CH3 )2S; this flux is thought by some to be as important as
or more important than H2S emissions from some wetlands. The general consensus,
however, is that most DMS comes from oceans as a product of decomposing phyto-
plankton cells and that the most important loss of sulfur from terrestrial freshwater
wetland systems is H2S.

Sulfide Oxidation

Sulfides can be oxidized by both chemoautotrophic and photosynthetic microorgan-
isms to elemental sulfur and sulfates in the aerobic zones of some wetland soils. Certain
species of Thiobacillus — and other bacteria collectively referred to as colorless sul-
fur bacteria (CSB) — obtain energy from the oxidation of hydrogen sulfide to sulfur,
whereas other species in this genus can further oxidize elemental sulfur to sulfate.
These reactions are summarized in Equations 6.10 and 6.11:

2H2S + O2 ^ 2S + 2H2O + (6.10)

and

2S + 3O2 + 2H2O ^ 2H2S4 + energy

(6.11)

The Carbon Cycle 195

Under anaerobic conditions, nitrate-nitrogen can be used as the terminal electron
acceptor in oxidizing hydrogen sulfides.

Photosynthetic sulfur-oxidizing bacteria, such as the green and purple sulfur bac-
teria found in salt marshes and mud fiats, are capable of producing organic matter in
the presence of light according to Equation 6.12:

CO2 + 2H2S + light ^ CH2O + 2S + H2O (6.12)

This reaction, called cmoxygenic photosynthesis, uses hydrogen sulfide as an electron
donor rather than H2O but is otherwise similar to the more traditional photosynthesis
equation. This reaction often takes place under anaerobic conditions where hydrogen
sulfide is abundant, but at the surface of sediments where sunlight is also available.

Sulfide Toxicity

Hydrogen sulfide, which is characteristic of anaerobic wetland sediments, can be toxic
to rooted higher plants and microbes, especially in saltwater wetlands where the con-
centration of sulfates is high. The negative effects of sulfides on higher plants include
the following:

1 . The direct toxicity of free sulfide as it comes in contact with plant roots;

2. The reduced availability of sulfur for plant growth because of its precipitation
with trace metals; and

3. The immobilization of zinc and copper by sulfide precipitation.

In wedand soils that contain high concentrations of ferrous iron (Fe^’*'), sulfides
can combine with iron to form insoluble ferrous sulfides (FeS), thus reducing the
toxicity of the free hydrogen sulfide. Ferrous sulfide gives the black color characteristic
of many anaerobic wetland soils; one of its common mineral forms is pyrite, FeS2, the
form of sulfur commonly found in coal deposits.

The Carbon Cycle

The major processes of carbon transformation under aerobic and anaerobic conditions
are shown in Figure 6.11. Photosynthesis (Equation 6.13) and aerobic respiration
(Equation 6.14) dominate the aerobic horizons (aerial and aerobic water and soil),
with H2O as the major electron donor in photosynthesis and oxygen as the terminal
electron acceptor in respiration:

6CO2 + I2H2O + light ^ C6H12O6 + 6O2 + 6H2O (6.13)

C5H12O6 + 6O2 ^ 6CO2 + 6H2O + \2e~ + energy (6.14)

The degradation of organic matter by aerobic respiration is fairly efficient in terms
of energy transfer. However, because of the anoxic nature of wedands, anaerobic

196 Chapter 6 Wetland Biogeochemistry

KEY

CH4

methane

C02

carbon dioxide

HC03

bicarbonate

carbonate

H,COt

carbonic acid

DOC

dissolved organic carbon

POC

particulate organic carbon

I Julflow

Figure 6.11 The carben cycle in wetiands. Majer pathways include photosynthesis, respira-
tion, fermentation, methanogenesis, and methane exidatien (anaerebic and aerobic). Also
indicated are the roies of sulfate and nitrate reduction in the carbon cycle.

processes, less efficient in terms of energy transfer, occur in proximity to aerobic pro-
cesses. Two of the major anaerobic processes are fermentation and methanogenesis.

Fermentation

The fermentation of organic matter, also called glycolysis for the substrate involved,
occurs when organic matter is the terminal electron acceptor in anaerobic respiration
by microorganisms and forms various low-molecular-weight acids and alcohols and
CO2. Examples are lactic acid (Eq. 6.15):

^ 2CH3CH2OCOOH (lactic acid) (6.15)

and ethanol (Eq. 6.16):

C6H12O6 ^ 2CH3CH2OH (ethanol) -I- 2CO2

(6.16)

The Carbon Cycle 197

Fermentation can be carried out in wetland soils by either facultative or obligate
anaerobes. Although in situ studies of fermentation in wedands are rare, fermentation
plays a central role in providing substrates for other anaerobes, such as methanogens
in wetland sediments. Fermentation represents one of the major ways in which
high-molecular-weight carbohydrates are broken down to low-molecular-weight
organic compounds, usually as dissolved organic carbon, which are, in turn, available
to other microbes.

Methanogenesis

Methano£ienesis occurs when certain bacteria ( methano£iens) use CO2 as an electron
acceptor for the production of gaseous methane (CH4), as described in Chapter 5:

CO2 + 8H+ ^ CH4 + 2H2O (6.17)

or, alternatively, use a low-molecular-weight organic compound, such as one from a
methyl group:

CH3COOH (acetic acid) ^ CH4 + CO2 (6.18)

3CH3OH (methanol) + 6H+ ^ 3CH4 + 3H2O (6.19)

Methane, which can be released to the atmosphere when sediments are disturbed,
is often referred to as sivampjjas or marsh ^as. Methane production requires extremely
reduced conditions, with a redox potential below -200 mV, after other terminal
electron acceptors (O2, NO3, and 804“) have been reduced. Methanogenesis is
carried out by methanopims — a group of microbes called the Archaea. Archaea are
prokaryotes that includes several obligate halophiles, and thermophiles in addition to
methanogens.

Methane Oxidation

Methane oxidation is carried out by obligate methanotropic bacteria^ which are from
a larger group of eubacteria; they convert methane gas in sequence to methanol
(CH3OH), formaldehyde (HCHO), and finally CO2:

CH4 ^ CH3OH ^ HCHO ^ HCOOH ^ CO2 (6.20)

Nonflooded lands (e.g., forests, agricultural land, grasslands) are normally consid-
ered the major biological sinks of methane and are where most methanotrophs occur.
But wetlands, which have stratified anoxic-oxic horizons, may have a lower anoxic
zone dominated by methanogenesis and a surface oxygenated zone with methane oxi-
dation (Fig. 6.11). Thus methane produced in the lower reaches of wetland soils may
be “modulated” by methanotrophs that intercept methane from below and convert it
to carbon dioxide. Methanotrophs are also able to tolerate extended periods of anoxia,
as with temporary flooding, and can resume methane oxidation within a few hours of

198 Chapter 6 Wetland Biogeochemistry

reexposure to oxygen (Whalen, 2005). Methanogens, however, are extremely sensi-
tive to oxygen; methane production does not continue very long once flooded soils
are drained. Roy-Chowdhury et al. (2014) found a high rate of potential methane oxi-
dation (PMO) by methanotrophs in Ohio created wetlands [equivalent to a methane
oxidation rate of 104 g-C m"^ yr M and also concluded that the soil methane concen-
tration had a greater influence than temperature on controlling methanotroph activity
in these wetlands.

In addition to methanotrophs, the autotrophic nitrifler communities discussed
previously are also able to carry out methane oxidation, because methane and ammo-
nia molecules have a similar size and structure. As a result, the ammonium molecule
can also essentially inhibit the methanotrophs from oxidizing CH4, and CH4 can
substitute for NH4“'‘ in nitriflers and be co-oxidized.

Methane Emissions

Methane emissions, which are the net result of methanogenesis and methane oxida-
tion, have a considerable range from both saltwater and freshwater wetlands as well as
from domestic wetlands, such as rice paddies. Comparison of rates of methane produc-
tion from different studies is difficult, because different methods are used and because
the rates depend on both soil temperature (season) and hydroperiod. Methane emis-
sions have clear seasonal patterns in temperate -zone wetlands (Fig. 6.12) and much
less seasonality in tropical and subtropical wetlands (Fig. 6.13). Summer rates can be
highest in seasonal climates, but estimation of total methane generation requires year-
long measurements, particularly in subtropical and tropical regions. The pattern also
depends on the degree of flooding and the presence or absence of vegetation. Stud-
ies of methane fluxes in temperate zone marshes have shown that methane fluxes are
higher in permanently flooded parts of the marshes than in intermittendy exposed
areas (Altor and Mitsch, 2006; Sha et al. 2011), suggesting that seasonal pulsing
rather than permanent flooding minimizes methane emissions. The lower rates of
methane generation in the intermittently exposed marshes could be a result of either
lower methanogenesis or higher rates of methane oxidation. Methane emissions in
tropical and subtropical climes show interesting patterns versus hydrologic conditions
(Fig. 6.14). In a series of tropical wetlands in different climates in Costa Rica, Nahlik
and Mitsch (2011) found a Shelford curve pattern with highest methane emissions
at middle water depths between 30 and 50cm (Fig. 6.14a). They attributed lower
methane emissions in shallow depths to better oxygen diffusion into the entire water
column, allowing oxidation of the soil-water interface and lower emissions in deepwa-
ter because of stratiflcation patterns typical in tropical bodies of water. Villa and Mitsch
(2014) describe a similar pattern for several plant communities in Corkscrew Swamp
area of the Greater Florida Everglades. Using a slightly different metric called days
after inundation (DAI) for the summer seasonal rains, they found that methane emis-
sions were low just after inundation, increased for about two months of inundation,
but began to decrease if flooding lasted longer than two months for a freshwater prairie
(Fig. 6.14b). This same pattern was also seen for several other wetland communities
at Corkscrew Swamp.

a. unvegetated

b. emergent vegetation

c. submerged vegetation

winter summer winter

Figure 6.12 Seasonal patterns of gas ebullition (flux of methane-rich bubbles) from three
different wetland community types in a floodplain lake (billabong) along the River Murray,
New South Wales, Australia: (a) no vegetation; (b) beds of the emergent plant Eleocharis
spbaceiata; and (c) beds of the submerged aquatic plant Vallisneria gigantea. Methane con-
centrations were 60 percent of the emissions from the bare area, 31 to 54 percent of the
emergent plant site, and 24 to 62 percent of the submerged aquatic plant site. (After Sorrell
and Boon, 1992)

199

200 Chapter 6 Wetland Biogeochemistry

Figure 6.13 Comparison of methane emission rates from Florida (subtropical) and Min-
nesota (temperate with cold winters), and model results that attempted to simulate both
conditions. (After Cui et al., 2005)

days after inundation pAI)

Figure 6.14 Relationships between hydrologic conditions and methane emissions for tropi-
cal and subtropical wetlands: (a) mean methane emissions versus water level for six wetland
transects in three tropical wetlands in Costa Rica (E = EARTH University campus wetland; LS
- La Selva Biological Station; PV - Palo Verde Biological Station); (b) methane emissions
(mean + standard error) versus days after inundation (DAI) for seasonally wet subtropical
wet prairie communities in Corkscrew Swamp Sanctuary in southwest Florida. Methane emis-
sions in bald cypress and pond cypress communities showed similar patterns versus DAI.
((a) Nahlik and Mitsch, 2011; (b) Villa and Mitsch, 2014)

Ebullition and Gaseous Transport in Plants

With the exception of CO 2 and O2, gases emitted from wetlands (1) emanate from
the sediment or soil surface through the water column by diffusive flux or diffusion,

The Carbon Cycle 201

(2) bubble to the surface in a process called ebuUitive flux or ebullition and then exit
to the atmosphere, or (3) pass through the vascular system of emergent plants (Boon,
1999). Boon and Sorrell (1995) noted that there were substantially more methane
fluxes during the day than during the night in chamber studies of Australian wetlands
when wetland plants were included in the chambers. They also noted that there was a
discrepancy in chambers between the total methane flux and the amount measured by
inverted funnels (which capture the ebuUitive flux). As a result, the pressures, flows,
and gas concentrations were measured within a dominant wetland plant, Eleocharis
sphacelata^ in both “influx” culms, which could generate high pressures, and “ef-
flux” culms, which could not. Methane concentrations were three orders of magnitude
greater in the efflux culms than in the influx culms. Carbon dioxide concentrations, as
expected, were 50 times higher, whereas dissolved oxygen concentrations decreased
20 percent. These studies and others suggest that between 50 and 90 percent of all
methane generated from a vegetated wetland could be passing through the vascular
system of emergent plants.

Carbon-Sulfur Interactions

The sulfur cycle is important in some wedands for the oxidation of organic carbon.
This is particularly true in most coastal wedands where sulfur is abundant. In general,
methane is emitted at low concentrations in reduced soils when sulfate concentradons
are high. Possible reasons for this phenomenon include (1) compeddon for substrates
that occurs between sulfur and methane bacteria, (2) the inhibitory effects of sulfate
or sulfide on methane bacteria, (3) a possible dependence of methane bacteria on
products of sulfur-reducing bacteria, and (4) a stable redox potential that does not
drop low enough to reduce CO 2 because of an ample supply of sulfate. Other evidence
suggests that methane may actually be oxidized to CO2 by sulfate reducers.

Sulfur-reducing bacteria require an organic substrate, generally of low molec-
ular weight, as a source of energy in converdng sulfate to sulfide (Eq. 6.9). The
process of fermentadon described previously can conveniendy supply these neces-
sary low-molecular-weight organic compounds, such as lactate or ethanol (see Eq.
6.15 and 6.16 and and Fig. 6.11). Equadons for sulfur reducdon, also showing the
oxidadon of organic matter, are shown in Equadons 6.21 and 6.22:

2CH3CHOHCOOH (lactate) -I- 804”^

^ 2CH3COOH -I- 2CO2+ H2S -I- 2H2O (6.21)

and

CH3COOH (acetate) -I- 804"^ ^ 2CO2 + H2S -I- 2H2O (6.22)

This fermentadon-sulfur reducdon pathway is pardcularly important in the oxida-
don of organic carbon to carbon dioxide in saltwater wetlands, which have an excess of
sulfates. Fully 54 percent of the carbon dioxide evolution from the salt marsh in New
England was caused by the fermentadon-sulfur reducdon pathway, with aerobic res-
piradon accounting for another 45 percent. By contrast, most of the carbon flux from

202 Chapter 6 Wetland Biogeochemistry

freshwater systems is through the methane-methane oxidation pathway. In a freshwa-
ter billabong in Australia, Boon, and Mitchell ( 1995 ) demonstrated that methanogen-
esis accounted for 30 to 50 percent of the total benthic carbon flux and that a major
portion of the carbon fixed by plants leaves the wetland via methanogenesis.

In general, the release of carbon by methane production is dominant in freshwater
wetlands, whereas oxidation of organic carbon by sulfate reduction is dominant in
saltwater wetlands.

The Phosphorus Cycle

Phosphorus (Fig. 6 . 1 5 ) is one of the most important limiting chemicals in ecosystems,
and wetlands are no exception. It is a major limiting nutrient in northern bogs, fresh-
water marshes, and southern deepwater swamps. In other wetlands, such as agricul-
tural wetlands and salt marshes, phosphorus is an important mineral, although it is not
considered a limiting factor because of its relative abundance and biochemical stability.

Figure 6.15 The phosphorus cycle in wetlands. Major pathways illustrated are
plant/microbial uptake, mineralization, adsorption/precipitation, sedimentation, and anaero-
bic release.

The Phosphorus Cycle 203

Phosphorus retention is considered one of the most important attributes of natural
and constructed wetlands, particularly those that receive nonpoint source pollution or
wastewater.

Phosphorus occurs as soluble and insoluble complexes in both organic and inor-
ganic forms in wetland soils. Inorganic forms include the ions P04^“, HP04^“, and
^2^04“ (collectively referred to as orthophosphates) with the predominant form
depending on pH. Phosphorus also has an affinity for calcium, iron, and aluminum,
forming complexes with those elements when they are readily available. Phosphorus
occurs in a sedimentary cycle rather than in gaseous cycles, such as the nitrogen, sul-
fur, and carbon cycles described earlier. At any one time, a major proportion of the
phosphorus in wedands is tied up in organic litter and peat and in inorganic sediments,
with the former dominating peatiands and the latter dominating mineral soil wetlands.

The analytical measure of biologically available orthophosphates is sometimes
called soluble reactive phosphorus (SRP), although the equivalence among SRP,
exchangeable phosphorus, and orthophosphate is not exact. However, it is often
used as indicators of the bioavailability of phosphorus. Dissolved organic phosphorus
(DOP) and insoluble forms of organic and inorganic phosphorus are generally not
biologically available until they are transformed into soluble inorganic forms.

Although phosphorus is not directly altered by changes in redox potential as are
nitrogen, iron, manganese, and sulfur, it is indirectly affected in soils and sediments by
its association with several elements, especially iron, that are so altered. Phosphorus is
rendered relatively unavailable to plants and microconsumers by:

1 . The precipitation of insoluble phosphates with ferric iron, calcium, and
aluminum under aerobic conditions;

2. The adsorption of phosphate onto clay particles, organic peat, and ferric and
aluminum hydroxides and oxides; and

3 . The binding of phosphorus in organic matter as a result of its incorporation
into the living biomass of bacteria, algae, and vascular macrophytes.

There are three general conclusions about the tendency of phosphorus to pre-
cipitate with selected ions: (1) Phosphorus is fixed as aluminum and iron phosphates
in acid soils; (2) phosphorus is bound by calcium and magnesium in alkaline soils;
and (3) phosphorus is most bioavailable at shghtly acidic to neutral pH (Reddy and
DeLaune, 2008). The precipitation of metal phosphates and the adsorption of phos-
phates onto ferric or aluminum hydroxides and oxides are believed to result from the
same chemical forces, namely, those involved in the forming of complex ions and salts.

Co-precipitation of Phosphorus

In many surface water wetlands, high algal productivity can pull CO2 out of the water,
shift the whole carbonate equilibrium, and drive the pH as high as 9 or 1 0 on a diurnal
basis. Under these conditions, co-precipitation of phosphorus as it adsorbs onto calcite
and precipitates as calcium phosphate can be significant, just as precipitation of calcium
carbonate is also accelerated. In a study of created marshes in central Ohio, calcite
and dolomite were found in significant concentrations m the algal mat biomass and

204 Chapter 6 Wetland Biogeochemistry

wetland sediments but not in the river inflow, indicating that the precipitated calcite
was produced within the wetlands in signiflcant amounts. Phosphorus co-precipitating
with calcite was up to 47 percent of the total phosphorus contained in the algal mat
in these wedands, suggesting that phosphorus co-precipitation essentially doubled
the phosphorus removal capability of the algal mat (Liptak, 2000). Wetlands with
high algal productivity thus have two major pathways for phosphorus removal: the
assimilation of phosphorus by algal cells and co-precipitation of phosphates caused by
high pH created by the algal water column productivity.

The Phosphorus Cycle

The sorption of phosphorus onto clay particles is important in aquatic ecosystems.
It is believed to involve both the chemical bonding of the negatively charged phos-
phates to the positively charged edges of the clay and the substitution of phosphates
for silicate in the clay matrix. This clay-phosphorus complex is particularly important
for many wedands, including riparian wetlands and coastal salt marshes, because a
considerable pordon of the phosphorus brought into these systems by flooding rivers
and ddes is brought in sorbed to clay pardcles. Thus, phosphorus cycling in many
mineral soil wetlands tends to follow the sediment pathways of sedimentadon and
resuspension. Because most wedand macrophytes obtain their phosphorus from the
soil, sedimentadon of phosphorus sorbed onto clay particles is an indirect way in which
the phosphorus is made available to the biotic components of the wetland. In essence,
the plants transform inorganic phosphorus to organic forms that are then stored in
organic peat, mineralized by microbial acdvity, or exported from the wedand.

The Phosphorus Cycle

When soils are flooded and condidons become anaerobic, several changes in the avail-
ability of phosphorus result. A well-documented phenomenon in the hypolimnion
of lakes is the increase in soluble phosphorus when the hypolimnion and the
sediment-water interface become anoxic. In general, a similar phenomenon often
occurs in wedands on a compressed verdcal scale. As ferric (Fe^'*‘) iron is reduced
to more soluble ferrous (Fe^“'‘) compounds, phosphorus that is in a specific ferric
phosphate (analydcally known as reductant-soluble phosphorus) is released into solu-
don. Other reacdons that may be important in releasing phosphorus upon flooding
are the hydrolysis of ferric and aluminum phosphates and the release of phosphorus
sorbed to clays and hydrous oxides by the exchange of anions. Phosphorus can also
be released from insoluble salts when the pH is changed either by the producdon
of organic acids or by the production of nitric and sulfuric acids by chemosynthetic
bacteria. Phosphorus sorption onto clay particles, however, is highest under acidic to
slighdy acidic condidons.

Water Chemistry

The inputs of materials to wetlands occur through geologic, biologic, and hydro-
logic pathways. The geologic input from weathering of parent rock, although poorly

Water Chemistry 205

understood, may be important in some wetiands. Biologic inputs include photosyn-
thetic uptake of carbon, nitrogen fixation, and biotic transport of materials by mobile
animals such as birds. Except for gaseous exchanges such as carbon fixation in photo-
synthesis and nitrogen fixation, however, elemental inputs to wetlands are generally
dominated by hydrologic inputs.

Oceans and Estuaries

Wetlands such as salt marshes and mangrove swamps are continually exchanging tidal
waters with adjacent estuaries and other coastal waters. The chemistry of these waters
differs considerably from rivers, streams, and lakes. Although estuaries are places where
rivers meet the sea, they are not simply places where seawater is diluted with fresh
water. Table 6. 1 contrasts the chemical makeup of average river water with the average
composition of seawater. The chemical characteristics of seawater are fairly constant
worldwide compared with the relatively wide range of river water chemistry. Total
salinity typically range from 33 to 37 parts per thousand (ppt). Although seawater
contains almost every element that can go into solution, 99.6 percent of the salinity
is accounted for by 11 ions. In addition to seawater dilution, estuarine waters can also
involve chemical reactions when sea and river waters meet, including the dissolution
of particulate substances, flocculation, chemical precipitation, biological assimilation
and mineralization, and adsorption and absorption of chemicals on and into particles
of clay, organic matter, and silt. In most estuaries and coastal wetlands, biologically
important chemicals such as nitrogen, phosphorus, silicon, and iron come from rivers,
whereas other important chemicals such as sodium, potassium, magnesium, sulfates,
and bicarbonates/carbonates come from ocean sources.

Table 6.1 Average chemical concentrations (mg/L) of ocean
water and river water

Chemical

Seawater

“Average” River

Na+

10,773

6.3

Mg2+

1,294

4.1

Ca2+

412

K+

399

2.3

ci-

19,340

7.8

SO42-

2,712

11.2

HCO37CO32-

142

58.4

4.5

0.01

1.4

0.1

<0.01

0.7

SiOo

<0.1->104

13.1

0-0.5

0.2

0-0.07

0.02

Particulate organic carbon

0.01-10

5-10

Dissolved organic carbon

1-5

10-20

206 Chapter 6 Wetland Biogeochemistry

Streams, Rivers, and Groundwater

As precipitation reaches the ground in a watershed, it infiltrates into the ground, passes
back to the atmosphere through evapotranspiration, or flows on the surface as runoff.
When enough runoff comes together, sometimes combined with groundwater flow,
in channelized streamflow, its mineral content is different from that of the original
precipitation. The “average” concentration of dissolved materials in the world’s rivers
is compared to seawater in Table 6.1. There is not, however, a typical water quality
for surface and subsurface streams and rivers as there is for seawater. Figure 6.16 illus-
trates the cumulative frequency of the ionic composition of freshwater streams and
rivers in the United States. It shows, for example, the average concentrations of the
many ions at the 50 percent line. Average NO3 concentrations are about Img/L,
whereas the average for Mg^”*" about 10 mg/L and the average total dissolved solids is
approximately 500 mg/L. The curves demonstrate the wide range over which these
chemicals are found in streams and rivers.

The variability in concentrations of chemicals in runoff and streamflow is caused
by five factors:

1. Groundwater influence. The chemical characteristics of streams and rivers
depend on the degree to which the water has previously come in contact with
underground formations and on the types of minerals present in those
formations. Soil and rock weathering, through dissolution and redox
reactions, provides major dissolved ions to waters that enter the ground.

Figure 6.16 Cumulative frequency curves fer cencentrations of various dissolved minerals
in surface waters. Horizontal dashed line indicates median concentrations, 90 percent cumu-
lative frequency indicates the 90th percentile concentration, 50 percent indicates the 50th
percentile concentration, and so on. (After Davis and DeWiest, 1966)

Water Chemistry 207

The dissolved materials in surface water can range from a few milligrams per
liter, found in precipitation, to 500 or even 1,000 mg/L. The ability of water
to dissolve mineral rock depends, in part, on its nature as a weak carbonic
acid. The rock being mineralized is also an important consideration. Minerals
such as limestone and dolomite yield high levels of dissolved ions, whereas
granite and sandstone formations are relatively resistant to dissolution.

2. Climate. Climate influences surface water quality through the balance of
precipitation and evapotranspiration. Arid regions tend to have higher
concentrations of salts in surface waters than do humid regions. Climate also
has a considerable influence on the type and extent of vegetation on the land,
and it therefore indirectly affects the physical, chemical, and biological
characteristics of soils and the degree to which soils are eroded and
transported to surface waters.

3 . Geographic effects. The amounts of dissolved and suspended materials that
enter streams, rivers, and wetlands also depend on the size of the watershed,
the steepness or slope of the landscape, the soil texture, and the variety of
topography. Surface waters that have high concentrations of suspended
(insoluble) materials caused by erosion are often relatively low in dissolved
substances. However, waters that have passed through groundwater systems
often have high concentrations of dissolved materials and low levels of
suspended materials. The presence of upstream wedands also influences the
quality of water entering downstream wetlands. Johnston et al. (2001) found
in a comparison of two riverine wedand areas of different soils and
geomorphology that there was nevertheless a seasonal convergence of surface
water chemistry caused by the wetlands that overrode the basin differences.

4. Streamflow/ ecosystem effects. The water quality of surface runoff, streams, and
rivers varies seasonally. There is generally an inverse correlation between
streamflow and concentrations of dissolved materials and streamflow. During
wet periods and storm events, the water is contributed primarily by recent
precipitadon that becomes streamflow very quickly without coming into
contact with soil and subsurface minerals. During low flow, some or much of
the streamflow originates as groundwater and has higher concentradons of
dissolved materials. The reladonship between pardculate matter and
streamflow is often the opposite. High flow often causes high concentradons
of sediments (pardculate matter).

5. Human effects. Water that has been modified by humans through, for
example, sewage effluent, urbanizadon, and runoff from farms often
drasdcally alters the chemical composidon of streamflow and groundwater
that reach wetlands. If drainage is from agricultural fields, higher
concentrations of sediments and nutrients and some herbicides and pesdcides
might be expected. Urban and suburban drainage is often lower than that
from farmland in those consdtuents, but it may have high concentrations of
trace organics, oxygen-demanding substances, and some toxins.

208 Chapter 6 Wetland Biogeochemistry

Nutrient Budgets of Wetlands

A quantitative description of the inputs, outputs, and internal cycling of materials in
an ecosystem is called an ecosystem mass balance. If the material being measured is one
of several elements such as phosphorus, nitrogen, or carbon that are essential for life,
then the mass balance is called a nutrient budget. In wetlands, mass balances have been
developed both to describe ecosystem function and to determine the importance of
wetlands as sources, sinks, and transformers of chemicals.

A general mass balance for a wetland, already shown in Figure 6.1, illustrates
the major categories of pathways and storages that are important in accounting for
materials passing into and out of wetlands. Nutrients or chemicals that are brought
into the system are called inputs or inflows. For wetlands, these inputs are primarily
through hydrologic pathways (described in Chapter 4), such as precipitation, surface
water and groundwater inflow, and tidal exchange. Biotic pathways of note that apply
to the carbon and nitrogen budgets are the fixation of atmospheric carbon through
photosynthesis and the capture of atmospheric nitrogen through nitrogen fixation.

Hydrologic exports., or losses or outflows, are by both surface water and ground-
water, unless the wetland is an isolated basin that has no outflow, such as a northern
ombrotrophic bog. The long-term burial of chemicals in the sediments is also consid-
ered a nutrient or chemical outflow, although the depth at which a chemical goes from
internal cycling to permanent burial is an uncertain threshold. The depth of available
chemicals is usually defined by the root zone of vegetation in the wetland. Biologically
mediated exports to the atmosphere are also important in the nitrogen cycle (denitri-
fication) and in the carbon cycle (respiratory loss of CO2). The significance of other
losses of elements to the atmosphere, such as ammonia volatilization and methane and
sulfide releases, are potentially important pathways for individual wetlands as well as
for the global cycling of minerals.

Intrasystem cycling involves exchanges among various pools, or standing stocks, of
chemicals within a wetland. This cycling includes pathways such as litter production,
remineralization, and various chemical transformations discussed earlier. The translo-
cation of nutrients from the roots through the stems and leaves of vegetation is another
important intrasystem process that results in the physical movement of chemicals
within a wetland.

Figure 6.17 illustrates in detail some of the major pathways and storages that
investigators should consider when developing nutrient mass balances for wetlands.
Few, if any, investigators have developed a complete mass balance for wetlands that
includes measurement of all of the pathways shown in the figure, but the diagram
remains a useful guide.

A phosphorus budget developed for an alluvial river swamp in southern Illinois
showed that 10 times more phosphorus was deposited with sediments during river
Hooding (3.6 g P m^^ yr than was returned from the swamp to the river during
the rest of the year (Fig. 6.18). Thus, the swamp was a sink for a significant amount
of phosphorus and sediments during that particular year of flooding, although the
percentage of retention was low (3-4.5 percent) because a very large volume of phos-
phorus passed over the swamp (80.2 g P yr ^) during flooding conditions.

Atmospheric Deposition

N fixation

Ammonia

volatilization

—s.

Denitrifi-

cation

Burial ^ 28

Inflow of groundwater

Figure 6.17 Model of major chemical storages and flows in a forested wetland. Storages: L,
above-ground shoots or leaves; T, stems, branches, perennial above-ground storage; R, roots
and rhizomes; W, surface water; D, litter and detritus; S, near-surface sediments; B, deep
sediments essentially removed from internal cycling. Flows: 1 and 2 are exchanges of dis-
solved and particulate matter with adjacent waters; 3-5 are nitrogen fixation in sediments,
rhizosphere microflora, and litter; 6 is denitrification; 7 and 8 are groundwater inputs; 9 and
10 are atmospheric inputs (e.g., precipitation); 11 and 12 are throughfall and stemflow; 13
is uptake by roots; 14 is foliar uptake from surface water; 15 is foliar uptake directly from
precipitation; 16 and 17 are translocation from roots through stem to leaves; 18 is litterfall;
19 and 20 are translocation of materials from leaves back to stems and roots; 21 is leach-
ing from leaves; 22 is death/decay of roots; 23 is incorporation of detritus into peat; 24 is
adsorption from water to detritus; 25 is release from detritus to water; 26 is volatilization
of ammonia; 27 is sediment-water exchange; and 28 is long-term burial of sediments. (After
Nixon and Lee, 1986)

209

210 Chapter 6 Wetland Biogeochemistry

Figure 6.18 Annual phosphorus budget for alluvial cypress swamp in southern Illinois.
(After Mitsch et al., 1979)

Detailed nitrogen and carbon budgets for created marshes in Ohio are illustrated
in Figure 6.19. Both of these budgets illustrate the importance of accurate hydro-
logic measurements. Each also shows significant nitrogen and carbon sequestration in
the wetland soils.

Generalizations about Nutrient Budgets in Wetiands

Chemical balances that have been developed for various wetlands are extremely vari-
able, but four generalizations have emerged from these studies:

1. Seasonal patterns of nutrient uptake and release are characteristic of many
wetlands. In temperate climates, retention of certain chemicals, such as
nutrients, is greatest during the growing season, primarily because of higher

Nutrient Budgets of Wetlands 211

Figure 6.19 Annual wetland nutrient budgets for (a) nitrogen and (b) carbon for two cre-
ated riparian wetlands in central Ohio. (Data are from Batson et al., 2012 and Waletzko and
Mitsch, 2013)

microbial activity in the water column and sediments and secondarily because
of greater macrophyte productivity. For example, in cold temperate climates,
distinct seasonal patterns of nitrate retention are evident in many cases, with
greater retention during the summer months when warmer temperatures
accelerate both denitrification microbial activity and algal and macrophyte
growth.

2. Wetlands are frequently coupled to adjacent ecosystems through chemical
exchanges that significantly affect both systems. Ecosystems upstream of
wetlands are often significant sources of chemicals to wetlands, whereas
downstream aquatic systems often benefit either from the ability of wetlands
to retain certain chemicals or from the export of organic materials.

212 Chapter 6 Wetland Biogeochemistry

3. Nutrient cycling in wetlands differs from both deepwater aquatic and
terrestrial ecosystem cyclinpi in temporal and spatial dimensions. More nutrients
are tied up in sediments and peat in wetlands than in most terrestrial systems,
and deepwater aquatic systems have autotrophic activity more dependent on
nutrients in the water column than on nutrients in the sediments.

4. Anthropogenic changes have led to considerable chanpies in chemical cyclinpi in
many wetlands. Although wetlands are quite resilient to many chemical
inputs, the capacity of wetlands to assimilate anthropogenic wastes from the
atmosphere or hydrosphere is not limitless.

Recommended Readings

Reddy, K. R, and R. D. DeLaune. 2008. Biopfeochemistry of Wetlands. Boca Raton,
FL: CRC Press.

Schlesinger, W. H., and E. S. Bernhardt. 2013. Biopieochemistry 3rd ed. Amsterdam,
Netherlands: Academic Press/Elsevier.

References

Altor, A. E., and W. J. Mitsch. 2006. Methane flux from created riparian marshes:
Relationship to intermittent versus continuous inundation and emergent macro-
phytes. Ecological Engineering 28: 224-234.

Batson, J., U. Mander, and W. J. Mitsch. 2012. Denitriflcation and a nitrogen budget
of created riparian wetlands. Journal of Environmental Quality 41: 2024-2032.

Boon, P. 1. 1999. Carbon cycling in Australian wetlands: The importance of methane.
Verhandlunpten Internationale Vereini£un£ fiir Limnoloptie 27: 1-14.

Boon, P. 1., and A. Mitchell. 1995. Methanogenesis in the sediments of an Aus-
tralian freshwater wetland: Comparison with aerobic decay and factors controlling
methanogenesis. EEMS Microbiology Ecology 18: 174-190.

Boon, P. 1., and B. K. Sorrell. 1995. Methane fluxes from an Australian floodplain
wetiand: The importance of emergent macrophytes. Journal of North American
Bentholopfical Society 14: 582-598.

Cui, J., C. Li, and C. Trettin. 2005. Analyzing the ecosystem carbon and hydrologic
characteristics of forested wedand using a biogeochemical process model. Global
Change Biology 11: 278-289.

David, M. B., C. G. Flint, G. F. Mclsaac, L. E. Gentry, M. K. Dolan, and G. F. Czapar.
2013. Biophysical and social barriers restrict water quality improvements in the
Mississippi River basin. Environmental Science & Technology 47: 11928—11929.

Davis, S. N., and R. J. M. DeWiest. 1966. Hydrogeology. John Wiley & Sons, New
York. 463 pp.

Doering, O. C., J. N. Galloway, T. L. Theis, V. Aneja, E. Boyer, K. G. Gassman,
E. B. Cowling, R. R. Dickerson, W. Herz, D. L. Hey, R. Kohn, J. S. Lighy,
W. Mitsch, W. Moomaw, A. Mosier, H. Paerl, B. Shaw, and P. Stacey. 2011.

References 213

Reactive Nitrogen in the United States: An Analysis of Inputs, Flows, Consequences,
and Management Options. U.S. Environmental Protection Agency Science Advi-
sory Board Integrated Nitrogen Committee, EPA-SAB-11-013, Washington,
DC, 139 pp.

Elder, D. V., B. D. Eyre, and L. Davison. 2008. The contribution of anammox and
denitrification to sediment N2 production in a surface flow constructed wetland.
Environ. Sci. Technol. 42(24): 9144-9150.

Galloway, J. N., J. D. Aber, J. W. Erisman, S. P. Seitzinger, R. W. Howarth, E. B.
Cowling, and B. J. Cosby. 2003. The nitrogen cascade. BioScience 53: 341-356.

Hernandez, M. E., and W. J. Mitsch. 2006. Influence of hydrologic pulses, flood-
ing frequency, and vegetation on nitrous oxide emissions from created riparian
marshes. Wetlands 26: 862-877.

Hernandez, M. E., and W. J. Mitsch. 2007. Denitrification in created riverine wet-
lands: Influence of hydrology and season. Ecological Engineering 30: 78-88.

Hey, D. L., and N. S. Philippi. 1995. Flood reduction through wetland restoration:
The Upper Mississippi River Basin as a case study. Restoration Ecology 3: 4-17.

Johnston, C. A., S. D. Bridgham, and J. P. Schubauer-Berigan. 2001. Nutrient dynam-
ics in relation to geomorphology of riverine wetlands. Soil Science of America
Journal 65: 557-577.

Laanbroek, H. J. 1990. Bacterial cycling of minerals that affect plant growth in water-
logged soils: A review. Aquatic Botany 38: 109-125.

Ligi, T., M. Truu, K. Oopkaup, H. Nolvak, U. Mander, W.J. Mitsch, and J. Truu.
2015. The genetic potential of N2 emission via denitrification and ANAMMOX
from the soils and sediments of a created riverine treatment wetland complex.
Ecolopfical Engineering doi.0rg/lO.lOl6/j.ecoleng. 2014.09.072

Liptak, M. A. 2000. Water column productivity, calcite precipitation, and phosphorus
dynamics in freshwater marshes. Ph.D. dissertation. The Ohio State University,
Columbus.

Megonigal, J. P, M. E. Hines, and P. T. Visscher. 2004. Anaerobic metabolism:
Linkages to trace gases and aerobic processes. In W. H. Schlesinger, ed., Bio-
^eochemistry . Elsevier-Pergamon, Oxford, UK, pp. 317M:24.

Mississippi River/Gulf of Mexico Watershed Nutrient Task Force. 2008. Gulf
Hypoxia Action Plan 2008 for Reducinjj, Mitijjatinjj, and Controlling Hypoxia
in the ^Northern Gulf of Mexico and Improving Water Quality in the Mississippi
River Basin. Mississippi River/Gulf of Mexico Watershed Nutrient Task Force,
Washington, DC.

Mitsch, W. J., C. L. Dorge, and J. R. Wiemhoff. 1979. Ecosystem dynamics: A
phosphorus budget of an alluvial cypress swamp in southern Illinois. Ecology 60:
1116-1124.

Mitsch, W. J., J. W. Day, Jr., J. W. Gilliam, P. M. Groffman, D. L. Hey, G. W. Ran-
dall, and N. Wang. 2001. Reducing nitrogen loading to the Gulf of Mexico from
the Mississippi River Basin: Strategies to counter a persistent ecological problem.
BioScience 51: 373-388.

214 Chapter 6 Wetland Biogeochemistry

Mitsch, W. J., J. W. Day, Jr., L. Zhang, and R. Lane. 2005. Nitrate-nitrogen retention
by wetlands in the Mississippi River Basin. Ecological En£iineerin^ 24: 267-278.

Mitsch, W. J., and J. W. Day, Jr. 2006. Restoration of wetlands in the Mississippi-Ohio-
Missouri (MOM) River Basin: Experience and needed research. Ecological Engi-
neering 26: 55-69.

Mitsch, W. J, L. Zhang, K. C. Stefanik, A. M. Nahlik, C. J. Anderson, B. Bernal,
M. Hernandez, and K. Song. 2012. Creating wetlands: Primary succession, water
quality changes, and self-design over 15 years. BioScience 62: 237-250.

Murphy, J. C., R. M. Hirsch, and L. A. Sprague, 2013. Nitrate in the Mississippi
River and its tributaries, 1980-2010 — An update: U.S. Geological Survey
Scientific Investigations Report 2013-5169, 31 p., http://pubs.usgs.gov/sir/
2013/5169/.

Nahlik, A. M. and W. J. Mitsch. 2011. Methane emissions from tropical freshwater
wetlands located in different climatic zones of Costa Rica. Global Change Biology
17: 1321-1334.

Nixon, S. W., and V. Lee. 1986. Wetlands and Water Quality. Technical Report
Y-86-2, U.S. Army Corps of Engineers Waterways Experiment Station,
Vicksburg, MS.

Reddy, K. R. and R. D. DeLaune. 2008. Biogeochemistry of Wetlands. CRC Press,
Boca Raton, EL, 774 pp.

Roy-Chowdhury, T., W. J. Mitsch, and R. P. Dick. 2014. Seasonal methanotrophy
across a hydrological gradient in a freshwater wetland. Ecological Engineering 72:
116-124.

Sha, C., W. J. Mitsch, U. Mander, J. Lu, J. Batson, L. Zhang, and W. He. 2011.
Methane emissions from freshwater riverine wetlands. Ecological Engineering 37:
16-24.

Singer, P. C., and W. Stumm. 1970. Acidic mine drainage: The rate-determining step.
Science 167: 1121-1123.

Song, K., M. E. Hernandez, J. A. Batson, and W. J. Mitsch. 2014. Long-term deni-
trification rates in created riverine wetlands and their relationship with environ-
mental factors. Ecological Engineering 72: 40-46.

Sorrell, B. K., and P. 1. Boon. 1992. Biogeochemistry of billabong sediments. 11.
Seasonal variations in methane production. Ereshwater Biology 27: 435-445.

Sundareshwar, P. V., J. T. Morris, E. K. Koepfler, and B. Fornwalt. 2005. Phosphorus
limitation of coastal ecosystem processes. Science 299: 563-565.

Villa, J. A. and W. J. Mitsch. 2014. Methane emissions from five wetland plant com-
munities with different hydroperiods in the Big Cypress Swamp region of Florida
Everglades. Ecohydrology & Hydrobiology 14: 253-266.

Waletzko, E. and W. J. Mitsch. 2013. The carbon balance of two riverine wet-
lands fifteen years after their creation. Wetlands 33: 989-999.

Whalen, S. C. 2005. Biogeochemistry of methane exchange between natural wedands
and the atmosphere. Environmental Engineering Science 22: 73-94.

Chapter ^

Wetland Vegetation and Succession

There are many plants (hydrophytes) adapted to temporary and permanent
floodinp! conditions found in wetlands. To counter anoxia, one important
structural adaptation in vascular plants is the development of pore space in the
cortical tissues, which allows oxygen to diffuse from the aerial parts of the plant
to the roots to supply root respiratory demands. There are many morphological
adaptations, such as pneumatophores, fluted trunks, prop roots, and
adventitious roots, that assist vascular plants in adapting! to havinp! their roots
in water. In addition, wetland plants have several physiological and whole
plant adaptations.

Wetland ecosystems have traditionally been considered transitional seres
between open lakes and terrestrial forests. The accumulation of organic
material from plant production was seen to build up the surface until it was no
lonpfer flooded and could support flood-tolerant terrestrial forest species
(autopfenic succession). An alternative theory is that the vepfetation found at a
wetland site consists of species adapted to the particular environmental
conditions of that site ( allopfenic succession). Current evidence seems to suggest
that both allogenic and autogenic forces act to change wetland vegetation.
Models used to describe wetland plant development include a functional guild
model, an environmental sieve model, and a centrifugal organization concept.

If one looks at ecosystem attributes as indices of succession, wetlands appear
to be mature in some respects and young in others. The strategy for ecosystem
development in wetlands includes concepts such as pulse stability and
self-organization or self-design. At landscape scales, patterns of wetlands,
aquatic and upland habitats reflect a complex and dynamic interaction of
physical (allogenic) and biotic (autogenic) forces.

215

216 Chapter 7 Wetland Vegetation and Succession

Wetland vegetation is generally viewed as consisting of vascular plants adapted to
flooding. Of the 325,000 vascular plant species in the world, only a small percent-
age have enough adaptations to be considered wetland plants. Wetland vegetation,
in the strict sense, also includes many unicellular species of algae and cyanobacteria.
Because of the significant number of metabolic and structural adaptations that wet-
land plants do have, there can be a wide diversity of plants in many wetlands. We often
see a minimum 100 vascular plant species in most mature wetlands, with perhaps half
of those listed as wetland plants. Defining what a wetland plant is exactly is difficult;
for example, it has remained one of the more challenging questions when wetlands
are defined legally in the United States.

Vascular Plant Adaptations to Waterlogging and Flooding

Wedand environments are characterized by stresses that most organisms are ill
equipped to handle. Aquatic organisms are not adapted to deal with the periodic
drying that occurs in many wetlands. Terrestrial organisms are stressed by long periods
of flooding. Because of the shallow water, temperature extremes on the wetland
surface are greater than would ordinarily be expected in aquatic environments. The
most severe stress, however, is the absence of oxygen in flooded wetland soils, which
prevents organisms from respiring through normal aerobic metabolic pathways. In
the absence of oxygen, the supply of nutrients available to plants is also modified, and
concentrations of certain elements and organic compounds can reach toxic levels.

Multicellular organization adds another layer of complexity to individuals com-
pared to unicellular organization. This complexity has enabled plants and animals
to develop a wider range of adaptations than bacteria to anoxia and to salt. At the
same time, some adaptations found in unicellular organisms, such as the ability to use
reduced inorganic compounds in the sediment as a source of energy, are not found in
multicellular organisms. These adaptations typically develop in specialized tissue and
organ systems.

In contrast to flood-sensitive plants, flood -tolerant species {hydrophytes) possess a
range of adaptations that enable them either to tolerate stresses or to avoid them. Sev-
eral adaptations by hydrophytes allow them to tolerate anoxia in wetland soils. These
adaptations can be grouped into three main categories: structural or morphological
adaptations, physiological adaptations, and whole plant strategies (Table 7.1). Details
of these adaptations are discussed here and throughout the chapter.

Morphological Adaptations

Aerenchyma

Virtually all hydrophytes have elaborate structural (or morphological) mechanisms to
avoid root anoxia. These responses to flooding are mechanisms that increase the oxy-
gen supply to the plant either by growth into aerobic environments or by enabling
oxygen to penetrate more freely into the anoxic zone. The primary plant strategy in

Vascular Plant Adaptations to Waterlogging and Flooding 217

Table 7.1 Plant adaptations and responses to flooding and
waterlogging

Structural (or Morphological) Adaptations

a. Aerenchyma tissue in roots and stem

b. Adventitious roots

c. Stem hypertrophy (e.g., buttress trunks)

d. Fluted trunks

e. Rapid vertical growth/grovrth dormancy

f. Shallow root systems/prop roots

g. Lentioles

h. Pneumatophores and cypress knees

Physiological Adaptations

a. Pressurized gas flow

b. Rhizospheric oxygenation

c. Decreased water uptake

d. Altered nutrient absorption

e. Sulfide avoidance

f. Anaerobic respiration

Whoie-Plant Strategies

a. Timing of seed production

b. Buoyant seeds and buoyant seedlings (viviparous seedlings)

c. Persistent seed banks

d. Resistant roots, tubers, and seeds

response to flooding is the development of air spaces {aerenchyma) in roots and stems,
which allow the diffusion of oxygen from the aerial portions of the plant into the roots
(Fig. 7.1). Aerenchyma development is not extensive in the absence of flooding and
is characteristic of flood -tolerant plant species, not flood -sensitive ones. In plants with
well-developed aerenchyma, the root cells no longer depend on the diffusion of oxy-
gen from the surrounding soil, the main source of root oxygen to terrestrial plants.
Unlike the plant porosity of normal plants, which is usually a low 2 to 7 percent of
volume, up to 60 percent of the volume of the roots of wetland species consists of
pore space. Air spaces are formed either by cell separation during maturation of the
root cortex or by cell breakdown. They result in a honeycomb structure. Air spaces
are not necessarily continuous throughout the stem and roots. The thin lateral cellu-
lar partitions within the aerenchyma, however, are not likely to impede internal gas
diffusion signiflcandy. The same kind of cell lysis and air space development has been
described in submerged stem tissue. Roots of flood-tolerant species, such as rice, form
aerenchyma even in aerated apical cells.

Root porosity is the overriding factor governing internal root oxygen con-
centration. The effectiveness of aerenchyma in supplying oxygen to the roots has
been demonstrated in several plant species. For example, the root respiration of
flood-tolerant Senecio aquaticuswzs only 50 percent inhibited by root anoxia, whereas
that of S. jacobaea, a flood-sensitive species, was almost completely inhibited. Greater

218 Chapter 7 Wetland Vegetation and Succession

(a) (b)

Figure 7.1 Light photomicrographs of Spartina alterniflora roots: (a) cross-section of a
streamside root; arrows indicate the presence of red ferric deposits on the root epidermis,
magnification xl92; (b) streamside root cross-section showing the presence of similar mate-
riais on the externai walis of the epidermal cells, magnification xl,143. Note the extensive
pore space (aerenchyma) in the roots. (From Mendelssohn and Postek, 1982)

root porosity in the tolerant species was the primary factor that contributed to the
difference. The most extensively studied flood-tolerant plant is rice. Rice plants grown
under continuous flooding develop greater root porosity than unflooded plants, and
this maintains the oxygen concentration in the root tissues. When deprived of oxygen,
rice root mitochondria degraded in the same way as did flood-sensitive pumpkin
plants, suggesting that the primary basis of resistance in flooded plants is by the avoid-
ance of root anoxia, not by physiological changes in cell metabolism (Levitt, 1980).

Adventitious Roots

In addition to aerenchyma development, anaerobic conditions result in the formation
of certain organs on wetland plants that assist the plant in getting oxygen to its root
system. Hormonal changes, especially the concentration of ethylene in hypoxic tissues,
initiate some of these structural adaptations. Ethylene has been reported to stimulate
the formation of adventitious roots in both flood-tolerant trees (e.g., Salix and Alnus)
and flood-tolerant herbaceous species (e.g., Phra^mites^ Ludwi^ia^ and Lythrum sali-
caria) and some flood-intolerant plants (e.g., tomato). These roots develop on the
stem just above the anaerobic zone when these plants are flooded (Fig. 7.2a). They
form as the original roots die and are able to function normally in an aerobic environ-
ment above the water line.

Stem Hypertrophy

Stem hypertrophy^ a noticeable swelling of the lower stem of vascular plants, is another
adaptation of many vascular plants to waterlogged conditions and hence serves as a
good indicator or wetland conditions. When this hypertrophy occurs on a tree, it is
called a buttress (Fig. 7.2b). It is a characteristic of swamp trees, such as bald and pond
cypress ( Taxodium spp. ) and water and swamp black gum ( Nyssa spp. ) . Hypertrophy is

Vascular Plant Adaptations to Waterlogging and Flooding 219

Figure 7.2 Illustrations of morphologicai adaptations to fiooding and waterlogging by vas-
cuiar plants: (a) adventitious roots on wiiiow {Salix) tree; (b) stem hypertrophy or buttresses
on cypress {Taxodium) trees in a deepwater swamp; (c) fluted trunk on pin oak tree {Quer-
cus palustris) in a freshwater forested wetland; (d) prop roots extending from Rhizophora
mangrove trees in Costa Rica; and (e) pneumatophores (“knees”) of Taxodium in a freshwa-
ter swamp. (Photo (a) by Ralph Tiner; (b), (c), (d), and (e) by W. J. Mitsch, reprinted with
permission)

not caused by the formation of aerenchyma but rather by larger cells and lower density
wood, also probably caused by ethylene production. A somewhat similar pattern of
trees exhibiting flared or fluted trunks (Fig. 7.2c) at the ground surface is common in
wetlands with several tree species, such as pin oak (Quercus palustris) and American
elm {Ulmus americana) .

220 Chapter 7 Wetland Vegetation and Succession

Stem Elongation, Root Adaptations, and Lenticels

Another response stimulated by submergence is rapid stem elongation in such aquatic
and semiaquatic plants as the floating heart {Nymphoides peltata), rice {Oryza sativa),
and bald cypress {Taxodium distichum), stimulated by rising water levels. Bald cypress
seedlings have rapid vertical growth rates supposedly to get the photosynthetic organs
out of harm’s way before standing water levels increase. The formation of shallow root
systems by wedand plants is another clear and common adaptation by vascular plants
to avoid anaerobic conditions. Deep taproots, common in upland forests, are almost
never found in forested wetlands. Some species are facultative m the regard. Red maple
[Acer rubrum) develops shallow root systems in wetlands but can have deep taproots
in upland forests.

The red mangrove {Rhizophora spp.) grows on arched prop roots in tropical and
subtropical tidal swamps around the world (Fig. 7. 2d). These prop roots have numer-
ous small pores, termed lenticels, above the tide level, which terminate in long, spongy,
air-filled, submerged roots. The oxygen concentration in these roots, embedded in
anoxic mud, may remain as high as 15 to 18 percent continuously, but if the lenticels
are blocked, this concentration can fall to 2 percent or less in two days. Lenticels are
also in the stems of flood-tolerant species, such as Alnusjjlutinosa and Nyssa sylvatica,
and serve as conduits to the aerenchymatous tissue in the stem.

Pneumatophores

Similarly, the black mangrove [Avicennia spp.) tree produces thousands of pneu-
matophores (air roots) about 20 to 30 cm high by 1cm in diameter, spongy, and
studded with lenticels. They protrude out of the mud from the main roots and are
exposed during low tides. The oxygen concentration of the submerged main roots
has a tidal pulse, rising during low tide and falling during submergence, reflecting
the cycle of emergence of the air roots. These pneumatophores are often covered
with lenticels that aid in root aeration. The “knees” of bald cypress {Taxodium
distichum) (Fig. 7.2e) are pneumatophores that improve gas exchange to the root
system. Cypress knees generally develop only when the trees are in waterlogged or
flooded soils, and their heights were often used as indicators of high water levels in
the wetlands.

Physiological Adaptations

Vascular emergent and floating-leaved wetland plants are sessile; only their roots are
in an anoxic environment. Typically, if the roots of a flood-sensitive upland plant
are inundated, the oxygen supply rapidly decreases. This shuts down the aerobic
metabolism of the roots, impairs the energy status of the cells, and reduces nearly
all metabolically mediated activities such as cell extension and division and nutrient
absorption. Even when cell metabolism shifts to anaerobic glycolysis, adenosine
triphosphate (ATP) production is reduced. Toxic metabolic end products of fermen-
tation may accumulate, causing cytoplasmic acidosis and eventually death. Anoxia is
soon followed by pathological changes in the mitochondrial structure. The complete

Vascular Plant Adaptations to Waterlogging and Flooding 221

destruction of mitochondria and other organelles occurs within 24 hours. Anoxia also
changes the chemical environment of the root, increasing the availability of reduced
forms of iron, manganese, and sulfur, which may accumulate to toxic levels in the
root. Several physiological adaptations of wetland vascular plants attempt to solve the
problem of anoxic conditions in the root system.

Pressurized Gas Flow

Dacey (1980, 1981) first described a particularly interesting adaptation that increases
the oxygen supply to the roots of the floating-leaved spatterdock ( Nuphar lutmm\ cur-
rently subdivided into several species including N. adventa). Since then, a similar adap-
tation of pressurized gas flow from the surface to the rhizosphere has been demon-
strated for other floating-leaved species. Fourteen emergent plants in southwestern
Australia were tested, and eight were found to have significant gas flow (0.2->10 cm^
min“^ culm”^), including Baumea articulata^Cyperus involucratus, Eleocharis sphace-
lata, Schoenopkctus validus^ Typha domingmsis, T. orimtalis^ Phm^mites australis^ and
Juncus ingms (Table 7.2). These results for such a wide variety of plants suggest that
internal pressurization and pressurized gas flow may be common to many hydrophytes.

Table 7.2 Pressurized gas flow in culms er leaves of 13 wetland plants and 1
upland plant in Australia^

Water Depth Species

AP^CPa)

Flow Rate (cm^ min culm ■'■)

Potentially Deepwater Plants

Phragmites australis

573 ± 54

5.3 ± 0.4"

Typha orientalis

1,070 ± 120

4.4 ± 0.3"

Typha domingensis

780 ± 140

3.4 ± 0.4"

Marginal Depth (<1 m) Plants

Juncus ingens

222 ± 24

1.2 ± 0.1"

Eleocharis sphacelata

1,080 ± 86

0.85 ± 0.02

Schoenoplectus validus

1,310 ± 124

0.29 ± 0.05

Baumea articulata

494 ± 58

0.23 ± 0.06

Very Shallow Water or Moist-Soil

Plants

Cyperus involucratus

903 ± 234

0.33 ± 0.09"

Canna sp.

27 ± 5

0.06 ± 0.01

Myriophyllum papillosum^

68 ± 12

0.04 ± 0.01

Cyperus eragrostis

111 ± 34

0.02 ± 0.01

Ludwigia pelloides'^

57 ± 1

<0.01

Bolboschoenus medianus

2 ± 31

<0.01

Not a True Wetland Plant

Arundo donax

1 ± 10

<0.01

“"Water depths refer to the potential depths that these plants can grow based on other studies and
plant size. AP^ refers to the static pressure differential in the plant stem. Plants are listed in order
of decreasing gas flow rates. Numbers indicate averages ± standard deviations.

'"Small specimens or leaves had to be removed to get flow rates within measuring range.

"Gas flow measured through detached culms.

“'Creeping, floating plants that grow in shallow water.

Source: Brix et al. (1992).

222 Chapter 7 Wetland Vegetation and Succession

Figure 7.3 [Gas flow.ai] Diel variation of soiar energy, temperature, internal pressure, and
gas flew for Typha domingensis: (a) photosynthetically available solar radiation (PAR), air
temperature (r^), internal leaf temperature (7j), and leaf surface temperature (7,); (b) inter-
nal static pressure differential (AP^) and gas flow. (From Brix et al., 1992)

T. domingensis had a dramatic diel pattern of convective gas flow related to air and
subsequent leaf temperatures (Fig. 7.3). Gas flows of 0.1 to 0.2 cm^ min“^ culm“^
occurred at night but increased to a rate as high as 3 cm^ min“^ culm“'^ during the
afternoon. The results were interpreted to suggest that humidity-induced pressuriza-
tion was the dominant driving force for the gas flow in the plants.

The pressure produced by this range of plants matched very nicely the approx-
imate depths at which these plants can potentially occur in wetlands (Table 7.2).
Phra^mites, Eleocharis, and the two Typha species can grow in water depths up to
2 m; Schoenoplectus, Juncus, and Ba-umea are found in water less than 1 m deep; the
two Cyperus species, Bolboschoenus^ and Canna^ grow in very shallow water to wet
soils. Air moves into the internal gas spaces (the lacunar system) of aerial leaves and is
forced down through the aerenchyma of the stem into the roots by a slight pressure
(~200-l,300 Pa) generated by a gradient in temperature and water vapor pressure.
Older leaves often lose their capacity to support pressure gradients, and so the return
flow of gas from the roots is through the older leaves, which are rich in carbon dioxide

Vascular Plant Adaptations to Waterlogging and Flooding 223

and methane from root respiration. The gas exchange to the rhizosphere through dead
culms of P. australis W2LS sufficient to maintain aerobic respiration of the plant roots.
The dead culms also provided an escape channel for excess CO2 and CH4 in the roots.

Grosse and others (Grosse and Schroder, 1984; Schroder, 1989; Grosse et al.,
1992) described a similar process in swamp trees, specifically in common alder (Alnus
jjlutinosa), the dominant tree species of European floodplain forests and riverine tem-
perate forests. Seedlings and dormant (leafless) trees of flood-tolerant species show
enhanced gas transport from aerial shoots to the roots when the shoots are heated by
the sun or incandescent light, compared to plants in the dark. Grosse et al. (1998)
called this phenomenon “pressurized gas flow” or “thermo-osmotic” gas flow. This
phenomenon occurs when a temperature gradient is established between the exterior
ambient air and the interior gas spaces in a plant’s cortical tissue. A second require-
ment is a permeable partition between the exterior and interior with pore diameters
“similar to or smaller than the ‘mean free path length’ of the gas molecules in the
system (e.g., 70 nm at room temperature and standard barometric pressure” (Grosse
et ah, 1998). In alder, meristematic tissue in the lenticels forms such a partition.

When the surface of the stem is mildly heated by sunlight, the mean free path
length of gas molecules in the intercellular spaces of the plant increases, preventing
the molecules from moving out through the osmotic barrier in the lenticels. The cooler
exterior molecules, however, can still diffuse into the plant. This sets up an internal
pressure gradient that forces gas down through the plant stem to the roots. This “ther-
mal pump” is not as effective in moving oxygen to the roots in alder as is extensive
aerenchyma tissue. For example, alder seedlings grown in flooded soil for two months
transported oxygen at eight times the rate of seedlings grown in aerated soil. The
difference was because of aerenchyma and lenticel development under flooded condi-
tions. By comparison, the thermo-osmotic effect (in the absence of flooding) led to a
fourfold increase in the rate of gas transport. The thermal pump is also not as active
in foliated trees as in dormant ones. Therefore, for trees, the adaptation appears to be
most effective in enhancing root aeration during seedling establishment in saturated
soils before aerenchyma development is accomplished, and in deciduous trees during
the dormant season.

Rhizosphere Oxygenation

Secondary effects of adaptations to root aeration influence other parts of the plants
or their environment. When anoxia is moderate, the magnitude of oxygen diffusion
through many wetland plants into the roots is apparently large enough not only to
supply the roots but also to diffuse out, oxidize the adjacent anoxic soil, and produce
an oxidized rhizosphere (see Ghapter 5: “Wetland Soils”). The brown deposits found
around the roots of Spartina alterniflora in Figure 7.1 were composed of iron and
manganese deposits formed when root oxygen comes in contact with reduced soil
ferrous ions. Oxygen diffusion from the roots is an important mechanism that mod-
erates the toxic effects of soluble reduced ions such as manganese in anoxic soil and
restores ion uptake and plant growth. These ions tend to be reoxidized and precipi-
tated in the rhizosphere, which effectively detoxifies them. In a similar vein, McKee

224 Chapter 7 Wetland Vegetation and Succession

et al. (1988) determined that soil redox potentials were higher and that pore water
sulfide concentrations were three to five times lower in the presence of the aerial prop
roots of the red mangrove (Rhizophom) or the pneumatophores of the black mangrove
{Avicennia) than in nearby bare mud soils, in all probability because of the diffusion
of oxygen from the mangrove roots into the soil. An interesting possibility is that the
root systems of these flood-tolerant plants may modify sediment anoxia enough to
allow the survival of nearby nontolerant plants (Ernst, 1990).

The presence of oxidized rhizospheres (now called oxidized pore linings by soil sci-
entists; see Chapter 5), which form as a result of root oxidation, is an important way in
which wetiands can be identified. Long after the plant roots die, residual veins of red
and orange, resulting from oxidized iron (Fe^'*‘ ) deposits, remain in many mineral soils,
a telltale sign that hydrophytes had been living in the soil. They are used in wetland
delineation practices as one indicator that hydric soils and, thus, wetlands are present.

Lower Water Uptake

Plants intolerant to anaerobic environments typically show decreased water uptake
despite the abundance of water, probably as a response to an overall reduction of root
metabolism. Decreased water uptake results in symptoms similar to those seen under
drought conditions: closing of stomata, decreased carbon dioxide uptake, decreased
transpiration, and wilting. The adaptive advantage of these responses is probably the
same as for drought-stricken plants — to minimize water loss and accompanying dam-
age to the cytoplasm. An accompanying depression of the photosynthetic machinery
is generally seen as an unavoidable corollary.

Sulfide Avoidance

Sulfur as sulfide is toxic to plant tissues. The element is reduced to sulfide in anaerobic
soils and accumulates to toxic concentrations, especially in coastal wedands. Although
sulfate uptake is metabolically controlled, sulfide can enter the plant without control
and is found in elevated concentrations in many flood -adapted species under highly
reduced conditions. In experiments with Spartina alterniflora^ a salt marsh species,
and Pcmicum hemitomon^ a freshwater marsh species, Koch et al. (1990) reported that
the activity of alcohol dehydrogenase (ADH), the enzyme that catalyzes the terminal
step in alcohol fermentation, was significantly inhibited by hydrogen sulfide and that
this inhibition may help explain the physiological mechanism of sulfide phytotoxicity
often seen in salt marshes. Sulfur tolerance in wetland plants varies widely, probably
because of the variety of detoxification mechanisms available. These include the oxi-
dation of sulfide to sulfate through root aeration of the rhizosphere; the accumulation
of sulfate in the vacuole; the conversion to gaseous hydrogen sulfide, carbon disulfide,
and dimethylsulfide and their subsequent diffusive loss; and a metabolic tolerance to
elevated sulfide concentrations.

Anaerobic Respiration

Under conditions of oxygen deprivation, plant tissues respire anaerobically, as
described for bacterial cells. In most plants, pyruvate, the end product of glycolysis.

Vascular Plant Adaptations to Waterlogging and Flooding 225

Figure 7.4 Schematic ef metabolic respiration pathway in flood-toierant plants. Left side of
diagram is aerobic respiration; right side is anaerobic respiration (fermentation/glycolsis),
which yields pyruvate, acetaldehyde, ethanol, and fatty acids such as maiic acid.

ADH: alcohol dehydrogenase.

NAD: nicotinamide adenine dinucleotide.

NADP: NAD phosphate.

Subscripts refer to oxidized (ox) and reduced (red) forms.

is decarboxylated to acetaldehyde, which is reduced to ethanol (see right side of
Fig. 7.4). Both of these compounds are potentially toxic to root tissues. Flood -tolerant
plants often have adaptations to minimize this toxicity. For example, under anaerobic
conditions, S. alterniflom roots show much increased activity of ADH, the inducible
enzyme that catalyzes the reduction of acetaldehyde to ethanol. The increase in the
enzyme indicates a switch to anaerobic respiration, and it explains why acetaldehyde
does not accumulate in the root tissue. Ethanol does not accumulate either, although
its production is apparendy stimulated. It diffuses from rice roots during anaerobiosis,
thus preventing a toxic buildup, and the same probably occurs in other wetland
plants. Another metabolic strategy reduces the production of alcohol by shifting the
metabolism to accumulate nontoxic organic fatty acids instead (Fig. 7.4). At one
time, it was suggested that malic acid (malate) accumulation may be a characteristic
feature of wetland species. The accumulation of malate cannot easily be interpreted,
however, in part because malate is an intermediate in several metabolic pathways.

226 Chapter 7 Wetland Vegetation and Succession

The metabolic problem encountered by plants deprived of oxygen is the loss of
the electron acceptor that enables normal energy metabolism through ATP forma-
tion and use. The metabolic bottleneck in this process is often the electron- accepting
coenzyme nicotinamide adenine dinucleotide (NAD), which is reduced in the oxida-
tive steps of carbohydrate metabolism and then reoxidized in the mitochondria by
molecular oxygen to yield the biological energy currency ATP (see shaded boxes in
left side of Fig. 7.4). In the absence of oxygen, reduced NAD (NAD,.gjj) accumulates
and “jams” the metabolic system, blocking ATP generation. In the process of fermen-
tation, acetaldehyde replaces oxygen, reoxidizing reduced NAD. Malate acts in the
same way through the tricarboxylic acid cycle. Thus, glycolysis can occur as long as
NAD is reoxidized to the oxidized form, NAD^,;.

Whole Plant Strategies

Many plant species have evolved avoidance or escape strategies by life-history adapta-
tions. The five most common of these strategies are listed next:

1 . The timing of seed production in the non-flood season by either delayed or
accelerated flowering

2. The production of buoyant seeds that float until they lodge on high,
unflooded ground

3 . The germination of seeds while the fruit is still attached to the tree
(vivipary), as in the red mangrove

4. The production of a large, persistent seed bank

5. The production of tubers, roots, and seeds that can survive long periods of
submergence

In many riparian wetlands, flooding occurs primarily during the winter and early
spring, when trees are dormant and much less susceptible to anoxia than they are
during the active growing season. The viviparous seedUnjjs that germinate live in the
canopy of red mangrove (Rhizophora) trees fall in the water from the canopy after ger-
mination and are transported, sometimes great distances. (See Chapter 9: “Mangrove
Swamps”). The seedling rights itself to a vertical position and develops roots if the
water is shallow until it lodges in shallow sediments, allowing the seedling to then
grow to a tree.

The freshwater aquatic monocot Sa^ittaria latifolia has a similar way of distribut-
ing its seeds. (The plant is sometimes called duck potato because its seeds resemble
potatoes.) The seed floats through a wetland until it lodges in a shallow area or amid
other emergent macrophytes, after which it germinates.

Mutualism and Commensalism

The close interactions among members of an ecological community reflect the high
degree of adaptability of members of the community, not only to their physical

Wetland Succession 227

environment but also to their biological environment. This chapter has documented
adaptations to the physical wetland environment, but positive interactions among
organisms are also predicted to play a significant role in ecosystem dynamics, especially
in marginal or stressed environments such as wetlands. Two such possibilities of
reactions between wedand populations are mutualism^ when there are positive and
obligatory benefits to both populations, and commensalism^ where one population
benefits and the other does not have either a positive or negative effect (i.e., it is
neutral).

Documentations of these effects in wetland environments are relatively few. Many
appear to involve nutrients that are limiting in these environments. For example,
Grosse et al. (1990) reported a commensalism or mutualism between alder {Alnus)
trees and fungi. Increased levels of nitrogen fixation by the symbiotic fungus Frankia
alni presumably occurred because of thermal pumping of oxygen through and out of
the root system of common alder into the rhizosphere. Ellison et al. (1996) reported
a mutualistic interaction between root-lbuling sponges {Todania i£;nis and Haliclona
implexiformis) and the red mangroves {Rhizophora manpile) on which they grow. Fine,
adventitious mangrove rootlets ramify throughout the sponges. They absorb dissolved
ammonium from the sponges, which stimulates additional root growth. The sponges
also protect the roots from isopod attack. Mangrove roots, in turn, provide the only
hard substrate for sponges in this habitat, and they stimulate sponge growth by leaking
carbon.

These examples of the positive interactions among wetland species point to
an extremely interesting line of neglected research that may lead to important
new insights into the complexity of mutualistic adaptations in wetland ecosystems
and their importance not only to the organisms involved but also to the energetic
dynamics of the entire community.

Wetland Succession

Allogenic versus Autogenic Succession

The beginning and subsequent development of a plant community is characterized by
the initial conditions at the site and by subsequent events, including the availability
of viable seeds or other propagules, appropriate environmental conditions for germi-
nation and subsequent growth, and replacement by plants of the same or different
species as site conditions change in response to both abiotic and biotic factors. The
concept of succession (i.e., the replacement of plant species in an orderly sequence of
development), in particular, has exerted a strong influence on plant ecology for more
than a century. Ecological theories of plant succession were advanced by H. C. Cowles
in his classic work on plant succession based on the sequential exposure of sand dunes
on the southern and eastern shores of Lake Michigan (Cowles, 1899). In that study,
dunes left bare from a retreating Lake Michigan were shown through a series of suc-
cessional ecosystems over thousands of years to go in an orderly primary succession to
a climax beech-maple forest (see Case Study 1).

228 Chapter 7 Wetland Vegetation and Succession

Autogenic succession was further enunciated by Clements (1916) and applied
to wetlands by the English ecologist W. H. Pearsall in 1920 and by an American,
L. R. Wilson, in 1935. E. P. Odum (1969) adapted and extended the ideas of those
early ecologists to include ecosystem properties such as productivity, respiration, and
diversity. This classical use of the term succession involves three fundamental concepts:
(1) vegetation occurs in recognizable and characteristic communities^ (2) community
change through time is brought about by the biota (i.e., changes are aut0£enic)\ and
(3) changes are linear and directed towdccd a mature, stable climax ecosystem. Although
this concept of autogenic succession was a dominating paradigm of great importance
in terrestrial ecology, the concept has been challenged and altered for almost a century.
Gleason (1917) enunciated an individualistic hypothesis to explain the distribution
of plant species. His ideas have developed into the continuum concept, which holds
that the distribution of a species is governed by its response to its environment {allo-
genic succession) . Because each species responds differently to its environment, no two
occupy exactly the same zone. The observed invasion/replacement sequence is also
influenced by the chance occurrence of propagules at a site. The result is a contin-
uum of overlapping sets of species, each responding to subtly different environmental
cues. In this view, no communities exist in the sense used by Clements, and although
ecosystems change, there is little evidence that this is directed or that it leads to a
particular climax.

A key issue in discussions of ecosystem development is whether biota determine
their own future by modifying their own environment, or if the development of an
ecosystem is simply a response to the external environment. In the classical view of
succession, wetlands are considered transient stages in the hydrarch development of a
terrestrial forested climax community from a shallow lake (Fig. 7.5). In this view, lakes
and open water gradually fill in as organic material from dying plants accumulates and
minerals are carried in from upslope. At first, change is slow because the source of
organic material is single-celled plankton. When the lake becomes shallow enough
to support rooted aquatic plants, however, the pace of organic deposition increases.
Eventually, the water becomes shallow enough to support emergent marsh vegetation,
which continues to build a peat mat. Shrubs and small trees appear. They continue to
transform the site to a terrestrial one, not only by adding organic matter to the soil but
also by drying it through enhanced evapo transpiration. Eventually, a climax terrestrial
forest occupies the site (Fig. 7.5a). The important point in this description of hydrarch
succession is that most of the change is brought about by the plant community as
opposed to externally caused environmental changes. A second important feature,
shown in Figure 7.5b, is that the process can reverse if the environmental conditions,
particularly the hydrology, change.

How realistic is this concept of succession? It is certainly well documented that
forests do occur on the sites of former lakes, but the evidence that the successional
sequence leading to these forests was autogenic is not clear. Because peat building is
crucial to filling in a lake and its conversion to dry land, key questions involve the con-
ditions for peat accumulation and the limits of that accumulation. Peat underlies many
wetlands, often in beds 10 m or more deep. In coastal marshes, peat has accumulated at

Presumed Succession Sequence

(b)

Figure 7.5 Classical hydrarch succession of freshwater wetlands: (a) succession from a
pond to a terrestrial ferest at the edge of a pond, and (b) general succession to mineral soil
forested wetlands in glaciated regiens ef North America, ((b) after Gelet et al., 1993)

229

230 Chapter 7 Wetland Vegetation and Succession

rates varying from less than 1 to up to 15 mm/yr. Most of this accumulation seems to
be associated with rising sea levels (or submerging land). By contrast, northern inland
bogs accumulate peat at rates of 0.2 to 2 mm/yr. (See Chapter 12: “Peatlands.”)

In general, accumulation occurs only in anoxic sediment. When organic peats are
drained, they rapidly oxidize and subside, as farmers who cultivate drained marshes
have discovered. As the wedand surface accretes and approaches the water surface or
at least the upper limit of the saturated zone, peat accretion in excess of subsidence
must cease. It is difficult to see how this process can turn a wetland into a dry habitat
that can support terrestrial vegetation unless there is a change in hydrologic condi-
tions that lowers the water table. For example, Cushing (1963) used paleoecological
techniques to show that most of the peatlands in the Lake Agassiz plain (Minnesota
and south-central Canada) formed during the mid-Holocene (beginning about 4,000
years ago) during a moist climatic period when surface water levels rose about 4 m.

Wetlands are at the center of the dispute about the importance of autogenic ver-
sus allogenic processes because of their transitional nature. In addition to being seres,
wetlands are often described as being ecotones — that is, transitional spatial gradients
between adjacent aquatic and terrestrial environments. Thus, wetlands can be consid-
ered transitional in both space and time. As ecotones, wedands usually interact strongly
to varying (allogenic) forcing functions from both ends of the ecotone. These forces
may push a wetland toward its terrestrial neighbor if, for example, regional water levels
fall, or toward its aquatic neighbor if water levels rise.

Alternately, plant production of organic matter may raise the level of the wetland,
resulting in a drier environment in which different species succeed. Because these
environmental changes can be subtle, it is often difficult to determine whether the
observed ecosystem response is autogenic or allogenic. Without careful measurements,
the causes of the response are often obscure.

CASE STUDY 1: Revisiting the Lake Michigan Dunes

In the early twentieth century, H. C. Cowles (1899, 1901, 1911) and Victor
Shelford (1907, 1911, 1913) studied ponds of different ages in the Indiana
dunes region along the southern shore of Lake Michigan. The ponds were
thought to represent an autogenic successional sequence. Along this age gra-
dient, the young ponds were deep and dominated by aquatic vegetation. Older
ponds were shallower and supported emergent vascular plants along their bor-
ders. The oldest ponds were shallowest and contained the most “terrestrial”
vegetation. This sequence was interpreted as evidence of classical autogenic
succession.

Wilcox and Simonin (1987) and Jackson et al. (1988) revisited the Indi-
ana dunes ponds. Using modern quantitative methods of ordination, they
found the same progression of plant species from young to old ponds, sup-
porting the sequence observed by earlier workers. In addition to the current
vegetation, they also examined pollen and macrofossils in the sediments of

Wetland Succession 231

a 3,000-year-old pond to determine whether the sediments support the pre-
sumed successional sequence found in the modern-day chronosequences.
Pollen and macrofossil data older than 150 B.P (before the present) consisted
of a diverse assemblage of submersed, floating-leaved, and emergent macro-
phyte groups. The data indicated a major and rapid vegetation change after
150 B.P, which the authors attributed to post-European settlement, such as
railroad construction and forest clearing.

Macrotossils

(wetland/aquatic)

100 60 60 40 20

total sum of squares

years BP

Core

present

2,800

5,700

10,400

- 20-

- 40-

- 60-
- 80-
-100-
-120'
-140
-160-
-180-
-200-
-220-
-240-
-260
-280-
_300 -

Pollen

(upland)

I 1 1 1 1 1

200 400 600

total sum of squares

Figure 7.6 Stratigraphically constrained ciuster diagrams for macrofossii and poiien
data from an indiana dunes pond in northern indiana adjacent to Lake Michigan. Macro-
fossil zonation is based on the presence or absence of aquatic and wetland taxa;
pollen zonation is based on percentages of selected upland pollen types. Close indi-
vidual taxa indicate close occurrence in fossil record. Thus the macrofossii record
indicates three different groups of organisms. (After Singer et al., 1996)

To further evaluate the historical changes in the Indiana dunes ponds.
Singer et al. (1996) examined the sediment pollen and macrofossii record of
aquatic and emergent plants in one of the old Indiana dunes pond sites and

232 Chapter 7 Wetland Vegetation and Succession

compared it with the regional terrestrial pollen record (of airborne pollen found
in the same cores). The latter tracks long-term climate changes in the region.
If aquatic and emergent paleotaxonomic remains showed changes in species
dominance that mirrored the terrestrial pollen record, then the changes in the
ponds could be attributed to regional climate change rather than to autogenic
processes. From their 10,000-year record. Singer et al. (1996) determined
that historic changes in pond vegetation did correspond to regional climate
change (Fig. 7.6). Between 10,000 and 5,700 B.P, the sampled area was
a shallow lake; the regional climate was mesic (a pine/oak/elm terrestrial
assemblage). A rapid increase in oak and hickory pollen around 5,700 B.P
signaled a regional climate shift to a drier environment. At the same level in
the sediment record, the pond macrofossil record showed a rapid shift to a
peat-forming marsh environment. After about 3,000 B.P, modest increases in
beech and birch pollen suggested a trend toward a cooler, moister climate. The
concomitant pond vegetation remained dominated by emergents, but transi-
tions among several taxa suggest that water-level fluctuations and occasional
fires were characteristic of the period.

These studies, taken together, provide a fuller, more complex picture
of plant development than the autogenic succession process proposed in
Cowles’s and Shelford’s earlier studies. The picture that emerges is one of
an interaction between allogenic and autogenic processes, with allogenic
forces driving the development of the biotic system, but modified by autogenic
processes. Over the 10,000-year span of the fossil record, changes in the
plant assemblage correlated well with regional climate change. However,
during the same period, the lake was slowly filling with organic sediments,
first 100 cm of gyttja, characteristic of open freshwater systems; then 200 cm
of fibrous peat, characteristic of vascular aquatic plants. During the period of
a slow climate shift to a less xeric environment after approximately 3,000 B.P,
the pond environment remained a marsh, although the species assemblage
changed. This fits well with the idea that the organic sediments moderated
the climatic influence on the local water levels. Finally, during the modern
period after about 150 B.P, human activities, which probably altered water
levels locally, resulted in rapid vegetation changes.

The Community Concept and the Continuum Idea

The Indiana dunes ponds example just described in Case Study 1 is only a small part of
the extensive literature concerning questions about plant and ecosystem development
in wetlands. The idea of the community is particularly strong in wetland literature.
Historic names for different kinds of wetlands — marshes, swamps, carrs, fens, bogs,
reedswamps — often used with the name of a dominant plant {Sphagnum bog, leather-
leaf bog, cypress swamp) — signify our recognition of distinctive associations of plants
that are readily recognized and at least loosely comprise a community. One reason

Wetland Succession 233

these associations are so clearly identified is that zonation patterns in wetiands often
tend to be sharp, having abrupt boundaries that call attention to vegetation change
and, by implication, the uniqueness of each zone. The plant community is central to
the historic idea of succession because the mature climax resulting from succession
was presumed to be a predictable group of plant species, with each group dependent
on the regional climate.

The identification of a community is also, to some extent, a conceptual issue
that is confused by the scale of perception. Field techniques are adequate to describe
the vegetation in an area and its variability. However, its homogeneity — one index
of community — may depend on size. For example, Louisiana coastal marshes have
been classified into four zones, or communities, based on the dominant vegetation.
If the size of the sampling area is large enough, any sample within one of these zones
will always identify the same species. If smaller grids are used, however, differences
appear within a zone. The intermediate marsh zone is dominated on a broad scale by
S. patens^ but aerial imagery shows patterns of vegetation within the zone, and inten-
sive sampling and cluster analysis of the vegetation reveal five subassociations that are
characteristic of intermediate marshes. Is the intermediate marsh a community? Are
the subassociations communities? Or is the community concept a pragmatic device to
reduce the bewildering array of plants and possible habitats to a manageable number
of groups within which there are reasonable similarities of ecological structure and
function?

Supporters of the continuum concept would argue that the scale dependence of
plant associations illustrates that individual species are simply responding to subde
environmental cues, implying little, if anything, about communities, and that plant
zonation simply indicates an environmental gradient to which individual species are
responding. The reason zonation is so sharp in many wetlands, they argue, is that
environmental gradients are “ecologically” steep, and groups of species have fairly
similar tolerances that tend to group them on these gradients.

One major difference between classical community ecologists and proponents of
the continuum idea is the greater emphasis put on allogenic processes by the latter. In
some wetlands, abiotic environmental factors often seem to overwhelm biotic forces.
In coastal areas, plants can do little to change the tidal pulse of water and salt. Tidal
energy may be modified by vegetation as stems create friction that slows currents or as
dead organic matter accumulates and changes the surface elevation. These effects are
limited, however, by the overriding tides. These wetlands are often in dynamic equi-
librium with the abiotic forces, an equilibrium that is sometimes called pulse stability
(see discussion later in this chapter).

In the low-energy environment of northern peatlands, in contrast to tidal marshes,
hydrologic flows can be dramatically changed by biotic forces, resulting in distinctive
patterned landscapes. Thus, changes in wetlands may be autogenic but are not nec-
essarily directed toward a terrestrial climax. In fact, wedands in dynamically stable
environmental regimes seem to be extremely stable, contravening the central idea of
succession. Pollen profiles were used to determine the successional sequence in British
northern peatlands. Sequences were variable and there were reversals and skipped

234 Chapter 7 Wetland Vegetation and Succession

Figure 7.7 Successional sequences reconstructed from stratigraphic and palynologicai
studies of postglaciai British peatlands. Thicker lines indicate the more common transitions.
(After Walker, 1970)

stages that may have been influenced by the dominant species first reaching a site
(Fig. 7.7). A bog, not some type of terrestrial forest that hydrarch succession would
have predicted, was the most common endpoint in most of the sequences described.

Linear Directed Change

If plant species development on a site is determined by allogenic processes and is,
therefore, simply a response to environmental forcing processes, then the successional
concept of linear directed change makes little sense. Although the scientific literature
is replete with schematic diagrams showing the expected successional sequence from
wetland to terrestrial forest, most of these are based on observed zonation patterns (or
chronosequences), assuming that these spatial patterns presage the temporal pathway
of change.

However, paleological analyses of soil profiles (such as those discussed earlier for
the Indiana dunes ponds) provide the best evidence to evaluate the concept. These
records, mostly from northern peat bogs, suggest two generalizations: (1) In some
sites, the current vegetation has existed for several thousands of years; and (2) climatic
change and glaciation had major impacts on plant species composition and distribu-
tion; generally, bogs expanded during warm, wet periods and contracted during cool,
drier periods. Pollen sequences, however, are generally consistent across Europe and
North America, indicating a response to similar global climate shifts. West (1964),
as quoted in McIntosh (1985), wrote tellingly: “We may conclude that our present
plant communities have no long history in the Quaternary, but are merely tempo-
rary aggregations under given conditions of climate, other environmental factors, and
historical factors.”

Seed Banks

Seed banks, referred to as buried reserves of viable seeds (Keddy, 2010), are an impor-
tant component of wetland succession. Many studies have documented the role of
chance in the development of plant communities, especially in the early stages. The
chance development can be the result of the availability of a seed bank and a changing
environmental condition (e.g., the flooding of a site after years of dry conditions).

Wetland Succession 235

In this respect, studies of seed banks and their role in the introduction and invasion
of plant species have been important. If the development of plants on a site can be
explained only in terms of the response of individual species to local conditions, then
the previous history of the site is important because it determines what propagules
are present for future invasion. This — the sediment seed bank — has been found to be
extremely variable — both in space and in time. Pederson and Smith (1988) made these
five generalizations about freshwater marsh seed banks:

1. Marshes with drawdowns produce the greatest number of seeds.

2. Seed banks are dominated by the seeds of annual plants and flood-intolerant
species. Areas that contain emergent plants have greater seed densities than
mud flats. Perennials generally produce fewer seeds that have shorter viability
than annuals. They are more likely to reproduce by asexual means such as
rhizomes.

3. Seed distribution decreases exponentially with the depth of the sediment.

4. Water is a major factor in seed banks. Seeds are concentrated along drift lines.
The kinds of seeds produced depend on the flooding regime — by
submergents when deep flooded, by emergents when periodically flooded,
and by flood -intolerant annuals during drawdowns.

5. Saline zones produce few seeds. A salt marsh is an example of a perennial-
dominated system in which most reproduction is asexual.

The germination of seedlings from a seed bank is similarly influenced by many fac-
tors that vary in space and time. Environmental factors, such as flooding, temperature,
soil chemistry, soil organic content, pathogens, nutrients, and allelopathy, have been
shown to influence recruitment. Water, in particular, is a critical variable, because most
wetland plant seeds require moist but not flooded conditions for germination and early
seedling growth. As a result of this restrictive moisture requirement, it is common to
find even-aged stands of trees at low elevations in riparian wetlands, reflecting seed
germination during relatively uncommon years when water levels were unusually low
during the spring and summer.

Postrecruitment processes play a major role in the distribution of adult plants at a
site, leading to plant assemblages that cannot be predicted from the seed bank alone.
Thus, in coastal areas where the dominant plant, Spartina alterniflom^ occurs in large
monotypic stands, it is often the pioneer species and remains dominant throughout
the life of the marsh. In contrast, in tidal and nontidal freshwater marshes, the seed
bank is much larger and richer, and the first species to invade a site may later be
replaced by other species.

Models of Wetland Community Succession

Plant Species Functional Groups

Historically, although the community concept has been of immense value in ecology,
it has been criticized for being imprecise and not subject to accurate predictive models

236 Chapter 7 Wetland Vegetation and Succession

for ecological communities. Some ecologists have addressed this problem in different
ways. One approach is to describe vegetation communities in terms of guilds or the
more recent term functional groups thzt can be defined by measurable traits (Keddy,
2010). A functional group is defined as a group of functionally similar species in a
vegetation community. This approach has two advantages: (1) It collapses the large
number of vegetation species in a wetland to a manageable subset; and (2) species are
defined in terms of measurable functional properties. Boutin and Keddy (1993) illus-
trated a functional classification of 43 wetland plant species in eastern North America
according to 27 functional traits (Table 7.3). Figure 7.8 summarizes the results, which
groups the species according to their traits into three groups: (1) ruderal annuals, (2)
interstitial perennials, and (3) matrix perennials. These are further split into seven
guilds, ranging from obligate annuals that flowered in the first year and then die at
the end of the growing season to tall clonal dominant species with deep roots such
as Typha xglauca) that reproduce vegetatively with extensive lateral spread. Most of
the functional groups appear to fall along a continuum of life histories adjusted to
different light regimes, which is consistent with the results of other studies.

Environmental Sieve Model

Van der Valk’s (1981) environmental sieve model of wetland succession (Fig. 7.9) is
also a Gleasonian model and is similar to Keddy’s model in several ways. The presence
and the abundance of each species depend on its life history and its adaptation to

1 St year

high % flowering
no lateral spread
higher above ground
’ biomass

1st year
low % flowering
some lateral spread

compact growth form vigorous clonal spread
shallow jooting deeper rooting

many short stems few tall stems many shoots few shoots

high crown area small crown area low photosynthetic higher photosynthetic

obligate annual perennial

major functional groups

U OBLIGATE FACULTATIVE
ANNUALS ANNUALS

short tall tall

low % higher % deep rooting
flowering flowering zone

INTERSTITIAL CLONAL

PERENNIALS DOMINANTS

short
shallow rooting
zone

CLONAL

STRESS-

TOLERATORS

reed

clonal tussock

N '

( \

RUDERAL

INTERSTITIAL

MATRIX

Figure 7.8 Functional classification of 43 species of plants from various wetland habitats
in eastern North America, based on 27 plant traits dispiayed in Table 7.3. (After Boutin and
Keddy, 1993).

Wetland Succession 237

Table 7.3 Traits measured on wetland plant species for functional guild classification

A. Ttaits Measured on 1-Year-Old Plants in the Garden

10
11

16
17

18, 19

Life span:

1 = annuals

2 = facultative annuals (100% flowering)

3 = partly facultative annuals (>50<100 % flowering)

4 = perennials (<50% flowering)

Percentage flowering first year
Final height or highest height (cm)

Rate of shoot extension (cm/day):

log^height at day 94 - log„ height at day 36
Day 94 - Day 36

Total biomass at harvest (g)

Aboveground biomass (g)

Belowground biomass (g)

Ratio belowground/aboveground biomass
Photosynthetic area (cm^); includes leaves and green stems
Photosynthetic area/total biomass (cm^/g)

Photosynthetic area/total volume occupied by a plant (cm^/ml) measured by
displacement of water in graduated cylinder
Total biomass/total volume (g/ml)

Total number of tillers or shoots
Crown cover (cm^): {{D^ + D2>/4)^
where

= first measure of crown diameter
D2 = second measurement at right angle to first
Stem diameter at ground level (cm)

Depth to belowground system (cm)

Diameter of belowground system, i.e., rhizome or main roots (cm)

Shortest (18) and longest (19) distances between two shoots or tillers (measure of
degree of clumping of aerial stems) (cm)

B. Traits Measured on Plants in Natural Wetlands (Adult Traits)

20 Total height (cm)

21 Total number of tillers or shoots

22 Stem diameter at ground level (cm)

23, 24 Shortest (23) and longest (24) distances between two shoots or tillers (cm)

25 Diameter of belowground system, i.e., rhizome or main roots (cm)

26 Depth to belowground system (cm)

C. Ttait Measured under Greenhouse Conditions

27 Relative growth rate (RGR) (day ^) between days 10 and 30

Source: Boutin and Keddy (1993).

the environment of a site. In van der Valle’s model, all plant species are classified into
life-history types, based on potential life span, propagule longevity, and propagule
establishment requirements. Each life-history type has a unique set of characteristics
and, thus, potential behavior in response to controlling environmental factors such as
water-level changes. These environmental factors comprise the “environmental sieve”
in van der Valle’s model. As the environment changes, so does the sieve and, hence,
the species present.

238 Chapter 7 Wetland Vegetation and Succession

Dispersal

dependent

species

AD-I

AD-II

PD-I

PD-II

VD-I

VD-II

environmental sieve (state: drawdown)

wetland

vegetation

" "t" f f "r f 1 "

AS-I AS-II PS-1 PS-II VS-I
Seed bank species

VS-II

Potentially
extirpated
species
— ► AD-II

► AS-II

— ► PD-II
— ► PS-II

Key

potential life span propagate longevity

A - annual D - dispersal dependent species (short-lived seeds)

P - perennial with limited life span S - seed bank species (long-lived seeds)

V - vegetatively propagated perennial

propagule establishment requirement (e.g., hydrology)

I - species only established in absence of standing water

II - species can be established in standing water

Figure 7.9 General sieve model of Gleasonian wetland (freshwater marsh) succession
proposed by van der Valk (1981).

Centrifugal Organization Concept

Several other models of community change have been developed, although few have
been applied to wetlands. Grime (1979) proposed that changes in species compo-
sition and richness of herbaceous plants was related to the gradients of disturbance
and stress factors, which reduced biomass and determined which functional plant
strategies would work best. Tilman (1982) suggested that competition among plants
controlled community plant distribution, with each species limited by a different ratio
of resources and spatial heterogeneity of the resources.

Wisheu and Keddy (1992) combined aspects of both Grime’s and Tilman’s mod-
els to propose a model of centrifugal organization of plant communities (Fig. 7.10a).
Gentrifugal organization describes the distribution of species and vegetation types
along standing-crop gradients caused by combinations of environmental constraints.
Wisheu and Keddy (1992) summarize the concept as follows:

Gradients radiate outwards from a single core habitat to many different peripheral
habitats. The assumed mechanism is a competitive hierarchy where weaker
competitors are restricted to the peripheral end of the gradient as a result of a
trade-off between competitive ability and tolerance limits. The benign ends of the
gradients comprise a core habitat, which is dominated by the same species. At the
peripheral end of each axis, species with specific adaptations to particular sources of
adversity occur.

Wetland Succession 239

peripheral peripheral

habitat habitat

peripheral

habitat

peripheral

habitat

peripheral peripheral

habitat habitat

ice-scoured
river banks

ice-scoured _ _ _ _

sand and clay ^ Eieoc^unsspp

Sdrpus
Sparganum
Typha
^ -

sandy
river banks

sandy

shores

/ 7

. 2 a' E

' , ^ , ' \'lTiTI> swtales

\ ' ' ' ' ^ ^

V CV X 750 ' / '

'500

gravel

lakeshores

biomass, g/m^

beaver

ponds

Figure 7.10 Centrifugal organizatien medeis iiiustrating (a) transitions from core habitat
to peripheral habitats aiong resource or stress gradients (generai model) and (b) freshwa-
ter wetland pattern for eastern North America, where large, leafy species such as cattail
{Typha spp.) occupy the core habitat, while several different species and communities occupy
peripherai habitats stressed by infertile sand, ice scouring, and beaver activity. (From
Wisheu and Keddy, 1992).

The core habitat in wetlands has low disturbance and high fertility and is dom-
inated by species that form dense canopies, such as Typha in eastern North America
(Fig. 7.10b). Peripheral habitats represent different kinds and combinations of stresses
(infertility, disturbance) and support distinctive plant associations. The model allows
one to predict how changes in gradients and, hence, peripheral habitats will change
community composition. In the case of the Typha-corc centrifugal model shown in
Figure 7.10b, ice scouring, infertile sandy soils, flooding by beavers, and open shore-
lines are among the stresses that shift communities to less productive, albeit possibly

240 Chapter 7 Wetland Vegetation and Succession

more diverse, assemblages. In this model, rare species are restricted to the peripheral
habitats that may contain most of the biological diversity of a landscape, suggesting
that the model could be useful for the protection of rare and endangered plant species
and the conservation of biodiversity (Keddy, 2010).

So far in this chapter, we have discussed vegetation changes in wetlands. We sum-
marize this discussion with a statement by Bill Niering (1989):

Traditional successional concepts have limited useftilness when applied to wetland
dynamics. Wetlands typically remain wet over time exhibiting a wetland aspect rather
than succeeding to upland vegetation. Changes that occur may not necessarily be
directional or orderly and are often not predictable on the long term. Fluctuating
hydrologic conditions are the major factor controlling the vegetation pattern. The
role of allogenic factors, including chance and coincidence, must be given new
emphasis. Cyclic changes should be expected as water levels fluctuate. Catastrophic
events such as floods and droughts also play a significant role in both modifying yet
perpetuating these systems.

Ecosystem Development

E. P. Odum ( 1969) described the maturation of ecosystems as a whole (as distinct from
plants, communities and species) in an article entitled “The Strategy of Ecosystem
Development.” The concepts, in general, have withstood the test of time, and are
republished with some update in Odum and Barrett (2005), published three years
after Odum’s death. In ecosystem development, species composition in immature to
mature (climax) stages are less important than are ecosystem functions, such as those
described in Table 7.4. Immature ecosystems, Odum had observed, are characterized.

Table 7.4 Selected attributes for ecosystem development

Ecosystem

Type

Community Energetics

Community Structure

P:R

Ratio*

P:B

Ratio*

Net

Community

Production

Food

Chains

Total

Biomass and
Nonliving
Organic
Matter

Species

Diversity

Organism

Size

Developing

<1 or >1

High

Linear, grazing

Low

Increases

Small

Initially

Mature (Ciimax)

Low

Weblike, detrital

High

High or Declines

Large

Natural Selection

Biogeochemical Cycles

Regulation

Ecosystem Type

Growth Form

Life Cycle

Mineral Cycles

Internal Cycling

Resilience

Resistance

Developing

r-selection

Short, simple

Open

Not important

High

Low

Mature (Ciimax)

K-selection

Long, complex

Closed

Important

Low

High

*P = gross primary productivity; R = respiration; B = biomass
Sources'. E. P Odum (1969, 1971) and E. R Odum and Barrett (2005)

Wetland Succession 241

in general, by high production to biomass (P'.B) ratios; an excess of production over
community respiration {P:R ratio >1 ); simple, linear, grazing food chains; low species
diversity; small organisms; simple life cycles; and open mineral cycles. In contrast,
mature ecosystems such as old-growth forests, tend to use all of their production to
maintain themselves and therefore have P:R ratios about equal to 1 and little, if any, net
community production. Production may be lower than in immature systems, but the
quality is better; that is, plant production tends to be high in fruits, flowers, tubers,
and other materials that are rich in protein. Because of the large structural biomass
of trees in forested ecosystems, the P:B ratio is small. Food chains are elaborate and
detrital based, species diversity is high, space is well organized into many different
niches, organisms are larger than in immature systems, and life cycles tend to be long
and complex. Nutrient cycles are closed; nutrients are efficiently stored and recycled
within the ecosystem.

It is instructive to see how wetland ecosystems fit into this scheme of ecosystem
development. Do their ecosystem-level characteristics fit the classical view that all wet-
lands are immature transitional seres.^ Or do they resemble the mature features of a
terrestrial forest.^ Five conclusions can be made:

1. Wetland ecosystems have properties of both immature and mature ecosystems. For
example, nearly all of the nonforested wetlands have P-.B ratios intermediate
between developing and mature systems and PR ratios greater than I.
Primary production tends to be very high compared with most terrestrial
ecosystems. These attributes are characteristic of immature ecosystems.
However, all of the ecosystems are detrital based, with complex food webs
characteristic of mature systems.

2. The Odum model used live biomass as an index of structure or “information”
within an ecosystem. This relationship is reflected in the high P-.B ratios
(immature) of nonforested wetlands and the low P-.B ratios (mature) of
forested wetlands. In a real sense, however, peat should be considered a
structural element of wetlands because it is a primary autogenic factor
modifying the flooding characteristic of a wetland site. If peat is included in
biomass, herbaceous wetlands would have the high biomass and low P:B
ratios characteristic of more mature ecosystems. For example, a salt or fresh
marsh has a live peak biomass of less than 2kg/m^. However, the organic
content of a meter depth of peat (peats are often many meters deep) beneath
the surface is on the order of 45 kg/ nf . This is comparable to the
above-ground biomass of the most dense wetland or terrestrial forest. As a
structural attribute of a marsh, peat is an indication of a maturity far greater
than the live biomass alone would signify.

3. Mineral cycles vary widely in wetlands. They range from extremely open
riparian systems in which surface water (and nutrients) may be replaced
thousands of times each year to bogs in which nutrients are derived from
precipitation alone and are almost quantitatively retained. An open nutrient
cycle is a juvenile characteristic of wetlands, directly related to the large flux

242 Chapter 7 Wetland Vegetation and Succession

of water through these ecosystems. Even in a system as open as a salt marsh
that is flooded daily, however, about 80 percent of the nitrogen used by
vegetation during a year is recycled from mineralized organic material.

4. Spatial heterogeneity isptenerally well orp/anized in wetlands alon^ allogenic
gradients. The sharp, predictable zonation patterns and abundance of
land-water interfaces are examples of this spatial organization. In forested
wetlands, vertical heterogeneity is also well organized. This organization is an
index of mature ecosystems. In most terrestrial ecosystems, however, the
organization results from autogenic factors in ecosystem maturation. In
wetlands, most of the organization seems to result from allogenic processes,
specifically hydrologic and salinity gradients created by slight elevation
changes across a wetland. Thus, the “maturity” of a wetland’s spatial
organization consists of a high level of adaptation to prevailing microhabitat
differences.

5. Life cycles of wetland consumers are usually relatively short but are often
exceedingly complex. The short cycle is characteristic of immature systems,
although the complexity is a mature attribute. Once again, the complexity of
the life cycles of many wetland animals seems to be as much an adaptation to
the physical pattern of the environment as to the biotic forces. Many animals
use wedands only seasonally or only during certain life stages. For example,
small marsh fish and shellfish make daily excursions into wedands during high
ddes, redring to adjacent ponds during ebb tides. Many fish and shellfish
species migrate from the ocean to coastal wetlands to spawn or for use as a
nursery. Waterfowl use northern wedands to nest and southern wetlands to
overwinter, migradng thousands of miles between the two areas each year.

Strategy of Wetland Ecosystem Development

In the previous sections, we showed that wetlands possess attributes of both immature
and mature systems and that both allogenic and autogenic processes are important.
Allogenic processes are important as forcing functions, which include factors such as
hydrology and propagule introduction, change. Autogenic processes are important as
the biota begin to control some of the physics and chemistry, as illustrated in Figure
4.1. In this secdon, we suggest that in all wedand ecosystems there is a common
theme: Development insulates the ecosystem from its environment.

At the level of individual species, this occurs through genedc (structural and phys-
iological) adaptadons to anoxic condidons. At the ecosystem level, it occurs primarily
through peat producdon, which tends to stabilize the flooding regime and shifts the
main source of nutrients to recycled material within the ecosystem. In forests, shading
is important in regeneration following disturbance.

Turnover Rates and Nutrient Influxes

The intensity of water flow over and through a wedand can be described by the water
renewal rate the rado of throughflow to the volume stored on the site (see

Wetland Succession 243

water renewal rate
retimes per yr). with range

nitrogen loading. g*N m'^ yr"’

bog

inland

fresh

marsh

mangrove

tidal

fresh

marsh

salt

marsh

swamp

forest

riparian

forest

Figure 7.11 Renewal rates of water and nitrogen ioading of major wetland types. This figure
illustrates that wetlands have five orders of magnitude differences of hydrologic inflow and
nutrient inflow.

Chapter 4: “Wetland Hydrology”). In wetlands, varies by five orders of magni-
tude (Fig. 7.11), ranging from about 1 per year in northern bogs to almost 10,000
per year in swamp forests. The nutrient input to a wetland follows closely the water
renewal rate, because nutrients are carried to a site by water. The amount of nitrogen
delivered to a wetland site, for example, also varies by five orders of magnitude, rang-
ing from less than 1 g m”^ yr“^ in a northern bog to perhaps 10,000 g m“^ yr"^ in
a riparian forested wetland (Fig. 7.11). Of course, not all of this nitrogen is available
to plants in the ecosystem because, in many cases, it is fiowing through much faster

244 Chapter 7 Wetland Vegetation and Succession

than it can be immobilized, but these figures mdicate the potential nutrient supply to
the ecosystem.

Yet despite the extreme variability in these outside (allogenic) forces of hydrology
and nutrient inflows varying over 10,000 times, as illustrated in Figure 7.11, wetland
ecosystems are remarkably similar in many respects. Total stored biomass, including
peat to 1 m depth, ranges from 40 to 60 kg m~^ — ^less than twofold. Soil nitrogen
similarly varies only about threefold, from about 500 to 1,500 g m”^. Net primary
production (NPP), a key index of ecosystem function, ranges for wetlands from 400
(peat bogs) to 2,000 (forested wetlands) g m”^ yr~\ a factor of five.

Wetland Insularity

So allogenic forces vary by 10,000 times (five orders of magnitude) but functions of
wetlands vary by two to five times (not even one order of magnitude). Although some
studies of individual species (e.g., Spartina alterniflora) or ecosystems (e.g., cypress
swamps) have concluded that productivity is directly proportional to the water renewal
rate, when different wetland ecosystems that constitute greatly different water regimes
are compared, the relationship breaks down or at least is logarithmic. The apparent
contradiction may be explained primarily by the role of stored nutrients within the
ecosystem. As the large store of organic nutrients in the sediment mineralizes, it pro-
vides a steady source of inorganic fertilizer for plant growth. As a result, much of
the nutrient demand is satisfied by recycling, even in systems as open as salt marshes
and riparian wetlands. External nutrient inputs provide a subsidy to this basic supply.
Therefore, growth is often apparently limited by the mineralization rate, which, in
turn, is strongly temperature and hydroperiod dependent. Temperatures during the
growing season are uniform enough to provide a similar nitrogen supply to plants in
different wetland systems, except probably in northern bogs. There, low temperatures
and short growing seasons limit mineralization and restrict nutrient input. The combi-
nation of the two factors limits productivity. Thus, as wetland ecosystems develop, they
become increasingly insulated from the variability of the environment by storing nutri-
ents. Often the same process that stores nutrients (i.e., peat accumulation) also reduces
the variability of flooding, further stabilizing the system. The surface of marshes, in
general, is built up by the deposition of peats and waterborne inorganic sediments. As
the elevation increases, flooding becomes less frequent, and sediment input decreases.
In the absence of overriding factors, coastal wetland marshes in time reach a stable
elevation somewhere around local mean high water. The surfaces of riparian wetlands
similarly rise until they become flooded only infrequently. Northern bogs grow by peat
deposition above the water table, stabilizing at an elevation that maintains saturated
peat by capillarity. Prairie potholes may be exceptions to these generalizations. They
appear to be periodically “reset” by a combination of herbivore activity and long-term
precipitation cycles and to achieve stability only in some cyclic sense.

Pulse Stability

In contrast to the lack of evidence for community succession to a stable set of species,
which was summarized earlier, the concept of a progression toward a mature ecosystem

Wetland Succession 245

has greater merit. The attributes of a mature stable ecosystem place it in dynamic equi-
librium with its environment, and although individual species may come and go, a
mature ecosystem is stable in the sense that it has built-in mechanisms (species diver-
sity, nutrient storage, and recycling) that resist short-term environmental fluctuations.
In fact, the three Odums (W. E. Odum et ah, 1995) suggested that natural processes
pulse regularly, and the mature ecosystem responds in a pulsing steady state. In wet-
lands, examples of this phenomenon include salt marshes, tidal freshwater marshes,
riverine forests, and seasonally flooded freshwater marshes, all of which are functionally
similar despite marked differences in species composition, diversity, and community
structure. Odum et al. (1995) suggest that natural pulses such as tides pump energy
into ecosystems and enhance productivity. Biotic events are geared to and take advan-
tage of these pulses; for example, the influx of small fish into flooded marshes to feed
during high tide or the capturing of young fish in backwater oxbows and billabongs
during flooding, with the captured fish serving as food for wading birds during periods
of low water. This concept is referred to as pulse stability.

Self-Organization and Self-Design

Most wetland ecosystems are continually open to atmospheric, hydrologic, and biotic
inputs of propagules of plants, animals, and microbes. Self-organization., as discussed
by Howard T. Odum (1989), manifests itself in both microcosms and newly cre-
ated ecosystems, “showing that after the first period of competitive colonization, the
species prevailing are those that reinforce other species through nutrient cycles, aids
to reproduction, control of spatial diversity, population regulation, and other means.”
Self-organization is further defined as “the process whereby complex systems consist-
ing of many parts tend to organize to achieve some sort of stable, pulsing state in the
absence of external interference” (E. P. Odum and Barrett, 2005).

Self-design, defined as “the application of self-organization in the design of ecosys-
tems” (Mitsch and Jorgensen, 2004), relies on the self-organizing ability of ecosys-
tems; natural processes (e.g., wind, rivers, tides, biotic inputs) contribute to species
introduction; selection of those species that will dominate from this gene inflow is
then nature’s manifestation of ecosystem design (Mitsch and Wilson, 1996; Mitsch
and Jorgensen, 2004; Mitsch et al., 2012). In self-design, the presence and survival
of species resulting from the continuous introduction of them and their propagules
is the essence of the successional and functional development of an ecosystem. This
can be thought of as analogous to the continuous production of mutations neces-
sary for evolution to proceed. In the context of ecosystem restoration and creation,
self-design means that, if an ecosystem is open to allow “seeding,” through human or
natural means, of enough species’ propagules, then the system will optimize its design
by selecting for the assemblage of plants, microbes, and animals that is best adapted
to the existing conditions. It is an important process to be investigated, particularly in
view of the interest in restoring and creating wedands.

In contrast to the self-design approach, the wetland restoration approach that is
still used today involves the introduction of organisms (often plants), the survival of
which becomes the measure of success of the restoration. This has sometimes been

246 Chapter 7 Wetland Vegetation and Succession

referred to as the “designer wetland” approach (Mitsch, 1998; van der Valle, 1998).
This latter approach, while understandable because of the natural human tendency to
control events, may be less sustainable than an approach that relies more on nature
being involved in the design.

Case Study 2: Wetland Primary Succession — A Wetland Experiment in
Self-Design

In a 20-year, whole-ecosystem experiment in two created freshwater marsh
basins at the Wilma H. Schlermeier Olentangy River Wetland Research Park at
The Ohio State University in central Ohio, Mitsch et al. (1998, 2005a, 2005b,
2012, 2014) describe how 2,500 individual wetland plants representing 13
species were introduced to one 1-ha flowthrough wetland basin while an adja-
cent identical wetland basin remained an unplanted control, essentially testing
the self-design capabilities of nature with and without human intervention.
Both basins have had identical inflows of river water and hydroperiods from
1994 through 2012. These experimental wetlands have allowed simultaneous
long-term study of three different questions related to wetland development:
(1) How Important is wetland plant introduction on long-term ecosystem func-
tion? (2) How long does it take for hydric soils and other wetland features to
develop at a site where no hydric soils previously existed? and (3) What are
the long-term patterns of blogeochemical changes of flowthrough wetlands as
they develop from open ponds of water to vegetated, hydric-soil marshes?

For the first six years of the wetland experiment, 17 different biotic and abi-
otic functional indicators of wetland function were measured, and similarities
of the wetland basins were estimated from these indicators. Indices were in six
different categories, including macrophytes, algal communities, water quality
changes, nutrient changes, benthic invertebrate diversity, and bird use. After
only three years, there appeared to be a convergence of wetland function of
the planted and unplanted basins, with a 71 percent similarity between the two
basins after only one year of divergence. By year 3, over 50 species of macro-
phytes, 130 genera of algae, over 30 taxa of aquatic invertebrates, and dozens
of bird species found their way naturally to both wetlands to supplement the
13 Introduced plant species (Mitsch et al., 1998). This convergence in year 3
followed the second year (1995), which showed only a 12 percent similarity
in the wetlands, probably because the planted wetland had macrophytes, but
the unplanted wetland did not.

The pattern of vegetation succession in the two experimental wetlands
for 17 years Is shown in Figure 7.12a. Not surprisingly, the clonal dominant
Typha spp. began to dominate the naturally colonizing wetland (unplanted
basin) quickly because there was little competition there. But Typha had min-
imal cover in the planted wetland until about the seventh year (2000) of the
experiment. Typha was held at bay in the planted wetland by competition

Wetland Succession 247

Wetland 1-Typha
Wetland 2-Typha

10000

8000

6000

4000

2000

* ANPP different in two
wetlands (n = 0.05)

Figure 7.12 Patterns of emergent macrophyte vegetation structure and function in
two experimentai wetiands in Ohio for 17 years (1994-2010) after wetiands were cre-
ated and one of the wetiands (Wetiand 1) was pianted in May 1994. (a) Percentage
cover of Typha spp. and other emergent macrophytes in pianted Wetiand 1 and natu-
raliy coionizing Wetiand 2. (b) Above-ground net primary productivity (ANPP; average
± std error) of planted Wetland 1 and naturally colonizing Wetland 2. (Updated from
Mitsch et al., 2012)

by several planted species, most notably Sparganium eurycarpum (bur reed)
and Schoenoplectus tabernaemontani (soft-stemmed bulrush). A muskrat
“eat-out" In 2000-2001 eliminated most of the Typha and other macro-
phytes in both wetlands, but a seed bank of one of the planted plants — S.
tabernaemontani — caused an explosion in cover in 2002, especially in the
wetland where it was planted and a year after the eat-out. After that, a

248 Chapter 7 Wetland Vegetation and Succession

10-year pattern of a slowly developing domination of both wetlands by Typha
occurred until both wetlands had an average cover of about 40 percent
Typha in 2009-2010 (years 16 and 17 since the wetlands were created).
The pattern of above-ground net primary productivity (ANPP) (Fig. 7.12b)
showed four years early in the experiment (1998-2001) when the so-called
unplanted wetland had higher productivity than did the planted wetland, mostly
because of Typha dominance. Then productivity decreased in both wetlands

1994 1996 1998 2000 2002 2004 2006 2008 2010

Figure 7.13 (a) Emergent macrophyte community diversity and (2) cumulative organic
productivity in two created experimental wetlands in Ohio for 17 years (1994-2010).
Wetland 1 (Wl) was planted in May 1994; Wetland 2 (W2) remained an unplanted
control. Hydrologic conditions were identical over that period in the two wetlands.
(Updated from Mitsch et al., 2012)

Wetland Succession 249

because of the muskrat eat-out but recovered quicker in the planted wetland,
which had higher productivity in 2003 and 2005.

In general, the planted wetland maintained a higher spatial macrophyte
diversity throughout much the study until the end (Fig. 7.13a) while the natu-
rally colonizing wetland has been more susceptible to disturbances such as
muskrat herbivory and hydrologic pulses than has the more diverse planted
wetland. More details of the vegetation richness over the 17 years are pre-
sented in Chapter 18: “Wetland Creation and Restoration."

The naturally colonizing wetland had benthic invertebrate diversity and
amphibian populations similar to the planted wetland for several of the early
years and had greater accumulated productivity after 17 years (Fig. 7.13b)
because of the several years when Typha dominated the unplanted wetland.

The continual introduction of species, whether introduced through flooding
and other abiotic and biotic pathways, appeared to have a much longer-lasting
effect in development of these ecosystems than the few species of plants
that were introduced to one of the wetlands in the beginning. This 17-year
study of wetland development showed that the planting had some long-term
effect on functions such as nutrient retention (Mitsch et al., 2012, 2014; see
also Chapter 19: “Wetlands and Water Quality") but none on denitrification
(Hernandez and Mitsch, 2007; Song et al., 2014). But the planting did appear
to have had a significant effect on decreasing carbon accumulation in the soil
(Anderson and Mitsch, 2006; Mitsch et al., 2012; Mitsch et al., 2013) and
methane emissions (Altor and Mitsch, 2006; Nahlik and Mitsch, 2010; Sha
et al. 2011). After 17 years, the two created wetlands may have converged
on plant cover structure, but the residual effect of the years when they were
different In productivity may still be influencing some ecosystem functions.

Ecosystem Engineers

Another way to describe the importance of autogenic successional processes involves
the introduced term ecosystem engineer — a term that is used to describe organisms
that have dramatic and important effects on an ecosystem (Jones et al., 1997; Alper,
1998). [This concept, also discussed in Chapter 4, should not be confused with
the field of ecological engineering, described by Mitsch and Jorgensen (2004) and
Mitsch (2012)]. In wedands, examples of ecosystem engineers could be muskrats
and beavers, both of which can have dramatic effects on vegetation cover and
ecosystem hydrology in freshwater marshes. In these cases, the biota does show a
dramatic feedback to the many features of the weriand (e.g., water levels or vegetation
productivity). One could argue that these ecosystem engineers “set back” succession
to an earlier stage; an alternative argument is that they are part of the ecosystem
development and that their behavior and its effects should be both expected and
appreciated as normal ecosystem behavior.

250 Chapter 7 Wetland Vegetation and Succession

(b)

Figure 7.14 Landscape patterns in Leuisiana wetiands: (a) a physically controlled pattern
of vegetation in tidal creeks in a Leuisiana salt marsh; (b) a biologically controlled pattern
of vegetation caused by a muskrat eat-out in a brackish marsh. Nete the high density of
muskrat lodges in (b)

Landscape Patterns

Many large wetland landscapes develop predictable and often complex patterns of
aquatic, wetland, and terrestrial habitats or ecosystems. In high-energy environments,
these patterns appear to reflect abiotic forces, but they are largely controlled by biotic

References 251

processes in low-energy environments. At the high-energy end of the spectrum,
the microtopography and sediment characteristics of mature floodplains — complex
mosaics of river channels, natural levees, back swamps, abandoned first and second
terrace flats, and upland ridges — reflect the flooding pattern of the adjacent river.
The vegetation responds to the physical topography and sediments with typical
zonation patterns. Salt marshes similarly develop a characteristic pattern of tidal
creeks, creekside levees, and interior flats that determine the zonation pattern and
vigor of the vegetation (Fig. 7.14a).

At the low-energy end of the spectrum, the characteristic pattern of strings and
flarks stretching for miles across northern peatlands appears to be controlled primarily
by biotic processes. Similarly, in many freshwater marshes, herbivores can be major
actors in the development of landscape patterns (Fig. 7.14b). In actuality, both phys-
ical (climatic, topographic, hydrologic) and biotic (production rates, root binding,
herbivory, peat accumulation) processes combine in varying proportions and interact
to produce observed wetland landscape patterns.

Recommended Readings

Keddy, P. A. 2010. Wetland Ecolojjy: Principles and Conservation, 2nd ed. Cambridge,
UK: Cambridge University Press.

van der Valk, A. G. 2012. The Biology of Freshwater Wetlands, 2nd ed. Oxford, UK:
Oxford University Press.

References

Alper, J. 1998. Ecosystem “engineers” shape habitats for other species. Science 280:
1195-1196.

Altor, A. E., and W. J. Mitsch. 2006. Methane flux from created riparian marshes:
Relationship to intermittent versus continuous inundation and emergent macro-
phytes. Ecolopfical Engineering 2B\ 224-234.

Anderson, C. J., and W. J. Mitsch. 2006. Sediment, carbon, and nutrient accumulation
at two 10-year-old created riverine marshes. Wetlands 26: 779-792.

Bernal, B. and W. J. Mitsch. 2013. Carbon sequestration in two created riverine wet-
lands in the Midwestern United States. Journal of Environmental Quality 42:
1236-1244.

Boutin, C., and P. A. Keddy. 1993. A functional classification of wetland plants. Jour-
nal ofVepietation Science 4: 591-600.

Brix, H., B. K. Sorrell, and P. T. Orr. 1992. Internal pressurization and convective
gas flow in some emergent freshwater macrophytes. Limnology and Oceanojjraphy
37: 1420-1433.

Clements, F. E. 1916. Plant Succession. Publication 242. Carnegie Institution of
Washington. 512 pp.

252 Chapter 7 Wetland Vegetation and Succession

Cowles, H. C. 1899. The ecological relations of the vegetation on the sand dunes of
Lake Michigan. Botanical Gazette 27\ 95-117, 167-202, 281-308, 361-369.

Cowles, H. C. 1901. The physiographic ecology of Chicago and vicinity. Botanical
Gazette 31: 73-108, 145-182.

Cowles, H. C. 1911. The causes of vegetative cycles. Botanical Gazette 51: 161-183.

Cushing, E. J. 1963. Late-Wisconsin pollen stratijjraphy in East-Central Minnesota.
Ph.D. dissertation. University of Minnesota, Minneapolis.

Dacey, J. W. H. 1980. Internal winds in water lilies: An adaption for life in anaerobic
sediments. Science 2\0: 1017-1019.

Dacey, J. W. H. 1981. Pressurized ventilation in the yellow waterlily. Ecology 62:
1137-1147.

Ellison, A. M., E. J. Farnsworth, and R. R. Twilley. 1996. Facultative mutualism
between red mangroves and root-fouling sponges in Belizean mangal. Ecology
77: 2431-2444.

Ernst, W. H. O. 1990. Ecophysiology of plants in waterlogged and flooded environ-
ments. Aquatic Botany 38: 73-90.

Gleason, H. A. 1917. The structure and development of the plant association. Torrey
Botanical Club Bulletin 44: 463M:81.

Golet, F. G., A. J. K. Galhoun, W. R. DeRagon, D. J. Lowry, and A. J. Gold. 1993.
Ecology of Red Maple Swamps in the Glaciated Northeast: A Community Profile.
Biological Report 12, U.S. Fish and Wildlife Service, Washington, DG. 151 pp.

Grime, H. H. 1979. Plant Stratejjies and Vegetation Processes. John Wiley & Sons,
New York.

Grosse, W., and P. Schroder. 1984. Oxygen supply of roots by gas transport in alder
trees. Zeitschrift fur Naturforsch. 39G: 1186-1188.

Grosse, W., S. Sika, and S. Lattermann. 1990. Oxygen supply of roots by
thermo-osmotic gas transport in Alnus-glutinosa and other wetland trees.
In D. Werner and P. Muller, eds. East Growinp; Trees and Nitrogen Fixing Trees.
Gustav Fischer Verlag, New York, pp. 246-249.

Grosse, W., J. Frye, and S. Lattermann. 1992. Root aeration in wetland trees by pres-
surized gas transport. Tree Physiology 10: 285-295.

Grosse, W., H. B. Biichel, and S. Lattermann. 1998. Root aeration in wetland trees
and its ecophysiological significance. In A. D. Laderman, ed., Coastally Restricted
Forests. Oxford University Press, New York, pp. 293-305.

Hernandez, M. E., and W. J. Mitsch. 2007. Denitrification in created riverine wet-
lands: Influence of hydrology and season. Ecological Engineering 30: 78-88.

Jackson, S. T, R. P. Rutyma, and D. A. Wilcox. 1988. A paleoecological test of a
classical hydrosere in the Lake Michigan dunes. Ecology 69: 928-936.

Jones, G. G., J. H. Lawton, and M. Shachak. 1997. Positive and negative effects of
organisms as physical ecosystem engineers. Ecology 78: 1946-1957.

Keddy, P. A. 2010. Wetland Ecology: Principles and Conservation, 2nd ed. Gambridge
University Press, Gambridge, UK. 497 pp.

References 253

Koch, M. S., I. A. Mendelssohn, and K. L. McKee. 1990. Mechanism for the hydro-
gen sulfide -induced growth limitation in wetland macrophytes. Limnology and
Oceanography 35: 399-408.

Levitt, J. 1980. Responses of Plants to Environmental Stresses, Voi. 11: Water, Radiation,
Salt, and Other Stresses. Academic Press, New York. 607 pp.

McIntosh, R. P. 1985. The Background of Ecology, Concept and Theory. Cambridge
University Press, Cambridge, UK. 383 pp.

McKee, K. L., 1. A. Mendelssohn, and M. W. Hester. 1988. Reexamination of
pore water sulfide concentrations and redox potentials near the aerial roots of
Rhizophora mangle and Avicennia germinans. American Journal of Botany 75:
1352-1359.

Mendelssohn, 1. A., and M. L. Postek. 1982. Elemental analysis of deposits on the
roots of Spartina alterniflora Loisel. American Journal of Botany 69: 904-912.

Mitsch, W. J. 1998. Self-design and wedand creation: Early results of a freshwa-
ter marsh experiment. In A. J. McComb and J. A. Davis, eds. Wetlands for the
Future. Contributions from INTECOL’s Fifth International Wetlands Confer-
ence. Gleneagles Publishing, Adelaide, Australia, pp. 635-655.

Mitsch, W. J. 2012. What is ecological engineering? Ecological Engineering 45: 5-12.

Mitsch, W. J., and R. F. Wilson. 1996. Improving the success of wetland creation
and restoration with know-how, time, and self-design. Ecological Applications 6:
77-83.

Mitsch, W. J., X. Wu, R. W. Nairn, P. E. Weihe, N. Wang, R. Deal, and C. E.
Boucher. 1998. Creating and restoring wetlands: A whole-ecosystem experiment
in self-design. BioScience 48\ 1019-1030.

Mitsch, W. J., and S. E. Jorgensen. 2004. Ecological Engineering and Ecosystem
Restoration. John Wiley & Sons, Hoboken, NJ. 411 pp.

Mitsch, W. J., N. Wang, L. Zhang, R. Deal, X. Wu, and A. Zuwerink. 2005a.
Using ecological indicators in a whole-ecosystem wetland experiment. In S. E.
Jorgensen, F-L. Xu, and R. Costanza, eds. Handbook of Ecological Indicators for
Assessment of Ecosystem Health. CRC Press, Boca Raton, EL, pp. 211-235.

Mitsch, W. J., L. Zhang, C. J. Anderson, A. Altor, and M. Hernandez. 2005b. Cre-
ating riverine wetlands: Ecological succession, nutrient retention, and pulsing
effects. Ecological Engineering 510-527.

Mitsch, W. J., B. Bernal, A. M. Nahlik, U. Mander, L. Zhang, C. J. Anderson,
S. E. Jorgensen, and H. Brix. 2013. Wetlands, carbon, and climate change.
Landscape Ecology 28\ 583-597.

Mitsch, W. J., L. Zhang, E. Waletzko, and B. Bernal. 2014. Validation of the ecosystem
services of created wetlands: Two decades of plant succession, nutrient retention,
and carbon sequestration in experimental riverine marshes. Ecological Engineering
72: 11-24.

254 Chapter 7 Wetland Vegetation and Succession

Nahlik, A. M., and W. J. Mitsch. 2010. Methane emissions from created riverine wet-
lands. Wetlands 30: 783-793 with Erratum in Wetlands {2Q11) 31: 449M:50.
Niering, W. A. 1989. Wetland vegetation development. In S. K. Majumdar,

R. R Brooks, F. J. Brenner, and J. R. W. Finer, eds. Wetlands Ecology and
Conservation: Emphasis in Pennsylvania. Pennsylvania Academy of Science,
Easton, pp. 103-113.

Odum, E. P. 1969. The strategy of ecosystem development. Science 164: 262-270.
Odum, E. P. 1971. Eundamentals of Ecology, 3rd ed. W. B. Saunders, Philadelphia.
544 pp.

Odum, E. P, and G. W. Barrett. 2005. Eundamentals ofEcolojjy, 5th ed. Thomson
Brooks/Cole, Belmont, CA, 598 pp.

Odum, H. T. 1989. Ecological engineering and self-organization. In W. J. Mitsch and

S. E. Jorgensen, eds. Ecological Engineering . John Wiley & Sons, New York, pp.
79-101.

Odum, W. E., E. P. Odum, and H. T. Odum. 1995. Nature’s pulsing paradigm. Estu-
aries 18: 547-555.

Pearsall, W. H. 1920. The aquatic vegetation of the English lakes. Journal of Ecology
8: 163-201.

Pederson, R. L., and L. M. Smith. 1988. Implications ofwedand seed bank research:
A review of Great Britain and prairie marsh studies. In D. A. Wilcox, ed.. Interdis-
ciplinary Approaches to Freshwater Wetlands Research. Michigan State University
Press, East Lansing, pp. 81-95.

Schroder, P. 1989. Characterization of a thermo-osmotic gas transport mechanism in
Alnus ^lutinosa (L.) Gaertn. Trees 3: 38M4.

Sha, C., W. J. Mitsch, U. Mander, J. Lu, J. Batson, L. Zhang and W. He. 2011.
Methane emissions from freshwater riverine wetlands. Ecological Enpfineerin^ 37:
16-24.

Shelford, V. E. 1907. Preliminary note on the distribution of the tiger beetle (Cicin-
dela) and its relation to plant succession. Biological Bulletin 14.

Shelford, V. E. 1911. Ecological succession. 11. Pond fishes. Biological Bulletin 21:
127-151.

Shelford, V. E. 1913. Animal Communities in Temperate America as Illustrated in the
Chicago Region. University of Chicago Press, Chicago.

Singer, D. K., S. T. Jackson, B. J. Madsen, and D. A. Wilcox. 1996. Differentiating
climatic and successional influences on long-term development of a marsh. Ecology
77: 1765-1778.

Tilman, D. 1982. Resource Competition and Community Structure. Princeton Univer-
sity Press, Princeton, NJ. 296 pp.

van der Valk, A. G. 1981. Succession in wetlands: A Gleasonian approach. Ecology 62:
688-696.

References 255

van der Valk, A. G. 1998. Succession theory and wetland restoration. In A. J. McComb
and J. A. Davis, eds., Wetlands for the Future. Contributions from FNTECOL’s Fifth
International Wetland Conference. Gleneagles Publishing, Adelaide, Australia, pp.
657-667.

Walker, D. 1970. Direction and rate in some British post-glacial hydroseres. In
D. Walker and R. G. West, eds. Studies in the Vc£ietational History of the British
Isles. Gambridge University Press, Gambridge, UK, pp. 117-139.

West, R. G. 1964. Inter-relations of ecology and quaternary paleobotany. Journal of
Ecology (Supplement) 52: 47-57.

Wilcox, D. A., and H. A. Simonin. 1987. A chronosequence of aquatic macrophyte
communities in dune ponds. Aquatic Botany 28: 227-242.

Wilson, L. R. 1935. Lake development and plant succession in Vilas Gounty, Wiscon-
sin. 1. The medium hard water lakes. Ecological Monographs 5: 207-247.

Wisheu, I. G., and P. A. Keddy. 1992. Gompetition and centrifugal organization of
plant communities: Theory and tests. Journal of Vegetation Science 3: 147-156.

Part III

Wetland Ecosystems

Tidal marshes: (a) tidal salt marsh in Louisiana; (h) tidal freshwater marsh in Maryland
(tidal freshwater photo courtesy of A. Baldwin)

Chapter 8

Tidal Marshes

The salt marsh, distributed worldwide alon£i coastlines in middle and hi£ih
latitudes, flourishes wherever the accumulation of sediments is equal to or
£ireater than the rate of land subsidence and where there is adequate protection
from destructive waves and storms. The important physical and chemical
variables that determine the structure and function of the salt marsh include
tidal floodinp! frequency and duration, soil salinity, soil permeability, and
nutrient limitation, particularly by nitropten. The vegetation of the salt marsh,
primarily salt-tolerant p/rasses and rushes, develops in identifiable zones in
response to these and possibly other factors. Mud and epiphytic alp/ae are also
often an important component of the autotrophic community. Heterotrophic
communities are dominated by detritalfood chains, with the p/razin^ food
chain beinp; much less sipfnificant except durinp; marsh die-off episodes.

Freshwater tidal marshes combine many features of both salt marshes and
inland marshes. They act in many ways like salt marshes, but the biota reflect
the increased diversity made possible by the reduction of the salt stress. Plant
diversity is hipfh, and more birds use these marshes than any other marsh type.
Because they are inland from the saline parts of the estuary, they are often close
to urban centers. This makes them more prone to human impact than coastal
salt marshes. Alonp! coastal rivers, tidal freshwater swamps tend to occupy a
narrow ran^e at the furthest extent of the tidal ran^e. They occur where tidal
waters are normally fresh and shallow enough for tree establishment.

Several types of wetlands in coastal areas are influenced by alternating floods and ebbs
of tides. Coastal wetlands include tidal salt marshes, tidal freshwater wetlands (marshes

259

260 Chapter 8 Tidal Marshes

Figure 8.1 Coastal wetlands lie on gradients of increasing saiinity from iniand to the ocean
in an estuary. Where saiinity is sufficient, sait marshes (in temperate zone) and mangroves
(in tropics) are found. Tidai freshwater marshes and tidal freshwater forests still experience
tides but are above the sait boundary. Farther inland are marshes and forested swamps that
experience neither salt nor tides.

and forests), and mangrove swamps. Salt marshes and tidal freshwater marshes are
discussed in this chapter. Mangrove swamps are discussed in the next chapter.

Near coastlines, the salinity of the water approaches that of the ocean (35 ppt),
whereas farther inland, the tidal effect can remain significant even when the salinity
is that of fresh water (Fig. 8.1). Tidal freshwater wedands are found upstream of salt
water (0.5 ppt = 500 ppm and lower salinity), while salt marshes and mangroves are
found downstream (salt marshes in temperate and boreal zones; mangroves in trop-
ics) in polyhaline and mesohaline estuarine waters greater than 5 ppt (=5,000 ppm) in
salinity. These coastal wetlands are found in abundance in the river deltas and estuar-
ies of the world — ^where large rivers debouch onto low-energy coasts (Fig. 8.2). These
river deltas and estuaries span the world’s latitudes and climatic zones. In the tropics,
tidally influenced wetlands of these deltas are mangroves. Above 25° latitude, man-
groves give way to salt marshes. In North America, large deltas are restricted to the
coasts of the South Atlantic and the Gulf of Mexico. The Mississippi River deltaic
marshes are the major example of this type of development and support the most
extensive coastal marshes in the United States.

We estimate that there are about 270,000 km^ of coastal wetlands in the world,
representing about 3 to 4 percent of all the wetlands in the world. About 150,000 km^
of those coastal wetlands are mangroves (see Chapter 9: “Mangrove Swamps”) with
tidal marshes (freshwater and salt) probably covering slighdy less than that amount
of area worlwide (Mitsch et ah, 2009). For the United States, the total area of

Tidal Salt Marshes 261

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Figure 8.2 The area of deltaic plains of selected major river systems of the world. (After
Coleman and Roberts, 1989)

wetlands considered coastal or estuarine wetlands, including Alaska, is approximately
32,000 km^, with about 19,000km^ as salt marsh, 8,000 km^ as tidal freshwater
marshes, and 5,000 km^ as mangrove swamps (Table 8.1). Almost 40 percent of the
salt marshes in continental United States are tbund in the Mississippi River Delta in
Louisiana (Ibanez et al., 2013).

Tidal Salt Marshes

Salt marshes are found throughout the world along protected coastlines in the middle
and high latitudes (Fig. 8.3a). Salt marshes can be narrow fringes on steep shore-
lines or expanses that are several kilometers wide. They are found near river mouths,
in bays, on protected coastal plains, and around protected lagoons. Different plant
associations dominate different coastlines, but the ecological structure and function

262 Chapter 8 Tidal Marshes

Table 8.1 Estimated area of coastal wetlands in the United States (x 1,000 ha)

Salt Marsh®

Freshwater Tidal Marsh"

Mangrove"

Total

Atlantic Coast

669

400

1,069

Gulf of Mexico

1,011

362

506

1,879

Pacific Coast

106

Alaska"

146

Total

1,875

819

506

3,200

®Watzin and Gosselink (1992)
"Field et al. (1991)

"Hall et al. (1994)

of salt marshes is similar around the world. Salt marshes, dominated by rooted veg-
etation that is alternately inundated and dewatered by the rise and fall of the tide,
appear from afar to be vast fields of grass of a single species. In reality, salt marshes
have a complex zonation and structure of plants, animals, and microbes, all tuned to
the stresses of salinity fluctuations, alternate drying and submergence, and extreme
daily and seasonal temperature variations. A maze of tidal creeks with plankton, fish,
nutrients, and fluctuating water levels crisscrosses the marsh, forming conduits for
energy and material exchange with the adjacent estuary. Studies of a number of differ-
ent salt marshes have found them to be highly productive and to support the spawning
and feeding habits of many marine organisms. Thus, salt marshes and tropical man-
grove swamps throughout the world form an important interface between terrestrial
and marine habitats.

Geographic Extent

Salt marshes are found near river mouths, in bays, on protected coastal plains, and
around protected lagoons. Different plant associations dominate different coastlines,
but the ecological structure and function of salt marshes is similar around the
world. Based the classification system developed by Valentine Chapman (1960,
1976), the world’s salt marshes can be divided into the following eight major
geographical groups:

1. Arctic. This group includes marshes of northern Canada, Alaska, Greenland,
Iceland, northern Scandinavia, and Russia. Probably the largest extent of marshes in
North America, as much as 300,000 km^ , occurs along the southern shore of the Hud-
son Bay. These marshes, influenced by ice, extreme low temperatures, a positive water
balance, and numerous inflowing streams, can be generally characterized as brack-
ish rather than saline. Various species of the sedge Carex and the grass Puccindlia
phry£ianodes often dominate. Parts of the southwestern coast of Alaska are dominated
by species of Salicornia and Suaeda.

2. Northern Europe. This group includes marshes along the west coast of Europe
from the Iberian Peninsula to Scandinavia, including Great Britain and the Baltic

Figure 8.3 Distribution of (a) sait marshes of the world and (b) wetiands in coastal drainage
areas of the United States, inciuding freshwater tidai wetlands and mangroves as well as
tidal salt marshes. (After Chapman, 1977 and Field et al., 1991)

263

264 Chapter 8 Tidal Marshes

Sea coast. Most of the western European coastal environment is characterized by a
moderate climate with sufficient precipitation but high salinities toward the southern
extremes. Dominant species include Puccindlm nmritima., Juncus ^emrdi, Salicor-
nia spp., Spartina anglica^ and S. townsendii. The west coast of Great Britain and
parts of the Scandinavian and Baltic Sea coasts, where substrates are dominated by
sand and salinities are low, are populated by Festuca rubra, A^rostis stolonifera, Carex
paleacea, Juncus bufonius, Desmoschomus bottanica, and Scripusspip. The muddy coast
of the English Channel is dominated by Spartina townsendii. Salt marshes in north-
ern Europe are often characterized by a lack of vegetation in the intertidal zone, in
contrast to North American marshes.

3. Mediterranean. This group includes the arid, rocky-to-sandy, high-salinity
coasts of the Mediterranean Sea. The salt marshes are dominated by low shrubby
vegetation, Arthrocnemum, Limonium,] uncus spp., and the halophyte Salicornia spp.

4. Eastern North America. These marshes, mostly dominated by Spartina and
Juncus species, are found along the eastern coasts of the United States and Canada
and the Gulf Coast of the United States. Salt marshes are most prevalent along the
eastern coast of the United States from Maine to Florida and on into Louisiana and
Texas along the Gulf of Mexico (Fig. 8.3b). The Eastern North American group is
further divided into three subgroups:

a. Bay of Fundy. River and tidal erosion is high in the soft rocks of this region,
producing an abundance of reddish silt. The tidal range, as exemplified at the
Bay of Fundy, is large, leading to a few marshes in protected areas and
considerable depth of deposited sediments. Puccinellia americana dominates
the lower marsh, and Juncus balticusis found on the highest levels.

b. New England. Marshes are built mainly on marine sediments and marsh peat,
and there is little transport of sediment from the hard-rock uplands. These
marshes range from Maine to New Jersey and are dominated by Spartina
alterniflora in the low marsh, with S. patens mixed with Distichlis spicata in
the high marsh.

c. Coastal Plain. These marshes extend southward from New Jersey along the
southeastern coast of the United States to Texas along the Gulf of Mexico.
Major rivers supply an abundance of silt from the recently elevated Coastal
Plain. The tidal range is relatively small. The marshes are laced with tidal
creeks. Mangrove swamps replace salt marshes along the southern tip of
Florida. Because of the extensive delta marshes built by the Mississippi River,
the Gulf Coast contains about 60 percent of the coastal salt and fresh marshes
of the United States (Fig. 8.3b). Dominant species are Spartina alterniflora,

S. patens, Juncus roemerianus, and Distichlis spicata.

5. Western North America. Compared with the Arctic and the eastern coast of
North America, salt marshes are far less developed along the western coasts of the
United States and Canada because of the geomorphology of the coastline. On this
rugged coast with its Mediterranean-type climate is found a narrow belt of Spartina

Tidal Salt Marshes 265

foliosa, often bordered by broad belts of Salicornm and Suaeda. Spartina alterniflora
is a nonnative invasive plant in coastal marshes north of California.

6. Australasia. Salt marshes are frequently found in river deltas along the tem-
perate coastlines of eastern Asia, Australia, and New Zealand on the Pacific Ocean,
Indian Ocean, and Tasman Sea.

a. Eastern Asia. The coasts of China, Japan, Russia, and Korea are generally
rugged and uplifted, with moderate precipitation but limited marsh
development. These marshes are dominated by Tripflochin maritima.,
Limonium japonicum., Salicornia., and Zoysia macrostachya. Major areas of salt
marsh rehabilitation have occurred on China’s eastern coastline, owing to the
introduction of Spartina an^Uca and S. alterniflora., although now both
species, especially S. alterniflora, are considered invasive and are being
eliminated where they compete with the native Phra^mites australis.

b. Australia. This group also includes New Zealand and Tasmania. It is
characterized by high rainfall and geographic isolation. Cosmopolitan species
in Australian salt marshes include Sporobolus virpfinicus, Sarcocornia
quinqueflora, and Suaeda australis. However, invasion of salt marshes of
Australia and New Zealand by several species is common. In New Zealand
and other temperate -zone salt marshes of the region, Spartina anpflica is a
major invasive species. Even with less rainfall and a clearly defined seasonal
pattern of wet and dry on the western coast of Australia, salt marshes can be
found, particularly around Shark Bay and the Peel-Harvey estuaries. In
contrast to the general case around much of the world, a majority of the salt
marshes of Australia are found in tropical regions (Adam, 1998).

7. South America. South American coasts too far south and too cold for man-
groves are rugged and geographically isolated. They are dominated by unique species
of Spartina, Limonium, Distichlis, Juncus, Heterostachys, and Allenrolfea.

8. Tropics. Although mangroves generally dominate tropical coastiines, salt
marshes are found in the tropics on high-salinity flats that mangroves cannot tolerate.
Spartina spp. and the halophytic genera Salicornia and Limonium often dominate.

Hydrogeomorphology

The physical features of tides, sediments, freshwater inputs, and shoreline structure
determine the development and extent of salt marsh wedands within their geograph-
ical range. Coastal salt marshes are predominandy interddal; that is, they are found
in areas at least occasionally inundated by high tide but not flooded during low tide.
A gende, rather than steep, shoreline slope allows for tidal flooding and the stability
of the vegetation. Adequate protecdon from wave and storm energy is also a physical
requirement for the development of salt marshes. Sediments that build salt marshes
originate from upland runoff, marine reworking of the coastal shelf sediments, or
organic producdon within the marsh itself

266 Chapter 8 Tidal Marshes

Table 8.2 Hydrologic demarcation between iow marsh and high marsh in
salt marshes

Marsh

Submergences
per Day in Daylight

Per Year

Maximum Period of
Continuous Exposure (days)

High marsh

<360

>10

Low marsh

>1.2

>360

Source: Chapman (1960)

Hydrology

Tidal energy represents a subsidy to the salt marsh that influences a wide range of
physiographic, chemical, and biological processes, including sediment deposition and
scouring, mineral and organic influx and efflux, flushing of toxins, and the control
of sediment redox potential. These physical factors in turn influence the species that
occur on the marsh and their productivity. The lower and upper limits of the marsh
are generally set by the tide range. The lower limit is set by the depth and the duration
of flooding and by the mechanical effects of waves, sediment availability, and erosional
forces. The upland side of the salt marsh generally extends to the limit of flooding on
extreme tides, normally between mean high water and extreme high water of spring
tides. Based on marsh elevation and flooding characteristics, the marsh is often divided
into two zones, the upper marsh {hijfh marsh) and the intertidal lower marsh {low
marsh) (Table 8.2). The high marsh is flooded irregularly and can experience at least
10 days of continuous exposure to the atmosphere, whereas the low marsh is flooded
almost daily, and there are never more than 9 continuous days of exposure. In the Gulf
Coast marshes of the United States, the terms streamside marshes and inland marshes
generally replace low and high marsh, respectively, because in these flat, expansive
marshes the streamside levees are actually the highest marsh elevations.

Marsh Development

Although a number of different patterns of development can be identifled, salt marshes
can be classifled broadly into two classes: (1) those that were formed from reworked
marine sediments on marine-dominated coasts; and (2) those that were formed in
deltaic areas where the main source of mineral sediment is riverine.

Marine-dominated marshes are typical of most of the world’s coastlines. On
marine-dominated coasts, salt marsh development requires sufficient shelter to
ensure sedimentation and to prevent excessive erosion from wave action. Marshes
can develop at the mouths of estuaries where sediments are deposited by the river,
behind spits and bars, and in bays that offer protection from waves and long-shore
currents. A spit is neck of land that acts to trap sediment on its lee side and protects
the marsh from the full forces of the open sea. The most extensive examples of this
type of coastal salt marsh in the United States have developed behind outer barrier
reefs along the Georgia-Carolina coast. Several large bays, such as Ghesapeake Bay,

Tidal Salt Marshes 267

Hudson Bay, the Bay of Fundy, and San Francisco Bay, are also protected adequately
from storms and waves so that they can support extensive salt marshes. These salt
marshes in bays have features of both marine and deltaic origins. They occur on the
shores of estuaries where shallow water and low gradients lead to river sediment
deposition in areas protected from destructive wave action. Tidal action must be
strong enough to maintain salinities above about 5 ppt; otherwise, the salt marsh will
be replaced by reeds, rushes, and other freshwater aquatic plants.

Major rivers carrying large sediment loads can build marshes in shallow estuaries
or out onto the shallow continental shelf where the ocean is fairly quiet. The size of a
delta increases with the size of the inflowing river’s drainage basin and its discharge,
but is modified by such factors as the slope of the ocean shelf into which the river
drains and the tidal range. In coasts with shallow slopes and low wave energies, deltas
can build out onto the shelf These deltas tend to have long shorelines relative to
their straight-line width. The interaction of river discharge and tidal energy determines
the salinity of the delta wetlands, with fresh river water reducing salinities and tidal
action extending the zone of marine -riverine interactions. One of the most dynamic
and expansive river-fed salt marshes is found in Mississippi River Delta in Louisiana.
Typically, the first marshes developing on newly deposited sediments are dominated by
freshwater species. However, the river course shifts through geologic time as the delta
lobe extends and the river loses efficiency. The abandoned marshes, no longer supplied
with fresh river water, become increasingly marine influenced. In the Mississippi River
Delta, these marshes undergo a 5,000-yr cycle of growth as fresh marshes, transition
to salt marshes, and finally degradation back to open water under the influence of
subsidence and marine transgression. During the last stage, the seaward edges of the
marshes are reworked into barrier islands and spits in the same way as marine -fed
coastal marshes on the Atlantic Coast.

Tidal Creeks

A notable physiographic feature of salt marshes, especially low marshes, is the devel-
opment of tidal creeks'm the marsh itself (Fig. 8.4). These creeks develop, as do rivers,
“with minor irregularities sooner or later causing the water to be deflected into defi-
nite channels” (Chapman, I960). The creeks serve as important conduits for material
and energy transfer between the marsh and its adjacent body of water. A tidal creek
has salinity similar to that of the adjacent estuary or bay, and its water depth varies with
tide fluctuations. Its microenvironments include different vegetation zones along its
banks that have aquatic food chains important to the adjacent estuaries. Because the
flow in tidal channels is bidirectional, the channels tend to remain fairly stable; that
is, they do not meander as much as streams that are subject to a unidirectional flow.
As marshes mature and sediment deposition increases elevation, however, tidal creeks
tend to fill in and their density decreases (Fig. 8.4).

Pannes

A distinctive feature of many salt marshes is the occurrence of pannes (pans). The
term panne is used to describe bare, exposed, or water-filled depressions in the marsh.

268 Chapter 8 Tidal Marshes

Figure 8.4 Drainage patterns of tidal creeks in young and mature Spartina alterniflora salt
marshes in the Dupiin River drainage, Doboy Sound, Georgia. (After Wiegert and Freeman,
1990, and Wadsworth, 1979)

which may have different sources. In the higher reaches of the marsh, inundated by
only the highest tides, sand barrens appear where evaporation concentrates salts in
the substrate, killing the rooted vegetation. These exposed barrens are often covered
by thin films of blue-green algae. Mud barrens are naturally occurring depressions in
the marsh that are intertidal and retain water even during low tide. Pannes are often
devoid of vascular vegetation or support submerged or floating vegetation because of
the continuous standing water and the elevated salinities when evaporation is high and
are continually forming and filling due to shifting sediments and organic production.
The vegetation that develops in a mud panne, for example, wigeon grass {Ruppia sp.),
is tolerant of salt at high concentrations in the soil water. Relatively permanent ponds
are formed on some high marshes and are flooded infrequendy by tides. Because of
their shallow depth and their support of submerged vegetation, they are used heavily
by migratory waterfowl. Pannes are a common feature due to human intervention,
occurring where free tidal movement has been blocked by roads or levees, where spoil
deposits have elevated a site, or where soil excavation, for example, for highway con-
struction, has occurred in a marsh.

Tidal Salt Marshes 269

Soil and Salinity

The sediment source and tidal current patterns determine the sediment characteristic
of the marsh. Salt marsh sediments can come from river silt, organic productivity in
the marsh itself, or reworked marine deposits. As a tidal creek rises out of its banks,
water flowing over the marsh slows and drops its coarser-grained sediment load near
the stream edge, creating a slightly elevated streamside levee. Finer sediments drop
out farther inland, giving rise to the well-known “streamside” effect, characterized
by the greater productivity of grasses along tidal channels than inland, a result of the
slightly larger nutrient input, higher elevation, and better drainage.

Salinity

Salt marshes that experience a large tide range (e.g., the Wash, England) tend to
approximate the ambient marine water salinity even though rainfall may be signifi-
cant. In coastal marshes adjacent to large rivers, in contrast (e.g., the north coast of
the Gulf of Mexico), fresh water dilutes marine sources, and the marshes are brack-
ish or even fresh. Extreme salinities can be found in subtropical areas, such as the
Texas Gulf Goast, where rivers and rainfall supply littie fresh water and tides have a
narrow range so that flushing is reduced. As a result, marine water is concentrated by
evapotranspiration, often to double seawater strength or even higher.

Lateral salinity gradients develop as a function of flooding frequency and sub-
sequently influence vegetation productivity (Fig. 8.5). Near the adjacent tidal creek,
frequent tidal inundation keeps sediment salinity at or below sea strength. As the marsh

Figure 8.5 The relatien of a salt flat's interstitial soil salinity and its vegetation. (After
Antifinger and Dunn, 1979 and Wiegert and Freeman, 1990)

270 Chapter 8 Tidal Marshes

elevation increases, the inundation frequency decreases and the finer sediments drain
poorly. At the salt fiat zone shown in Figure 8.5, infrequent spring tides bring in salt
water that is concentrated by evaporation. Flushing is not frequent enough to remove
these salts, so they accumulate to lethal levels. Above this elevation, tidal flooding is so
infrequent that salt input is restricted, and flushing by rainwater is sufficient to prevent
salt accumulation. In this way, the salt gradient set up by the interaction of marsh ele-
vation, tides, and rain often controls the general zonation pattern of vegetation and its
productivity. Within the salt marsh zone itself, however, all plants are salt tolerant, and
it is misleading to account for plant zonation and productivity on the basis of salin-
ity alone. Salinity, after all, is the net result of many hydrodynamic factors, including
slope and elevation, tides, rainfall, freshwater inputs, and groundwater. Thus, when
Spartitm flourishes in the intertidal zone, it is also responding to tides that reduce the
local salinity, remove toxic materials, supply nutrients, and modify soil anoxia. All of
these factors collectively contribute to different productivities and different growth
forms in the intertidal and high marshes.

Vegetation

The salt marsh ecosystem has diverse biological components, which include vegetation
and animal and microbe communities in the marsh and plankton, invertebrates, and
fish in the tidal creeks, pannes, and estuaries. The discussion here will be limited to
the biological structure of the marsh itself Plants and animals in these systems have
adapted to the stresses of salinity, periodic inundation, and extremes in temperature.

The vegetation of salt marshes can be divided into zones that are related to the
high and low marshes described previously but that also reflect regional differences.
Figure 8.6 shows a typical New England vegetation zonation pattern from streamside
to upland. The intertidal zone or low marsh next to the estuary, bay, or tidal creek is
dominated by the tall form of S. alterniflora Loisel (smooth cordgrass). In the high
marsh, S. alterniflom gives way to extensive stands of S. patens (saltmeadow cord-
grass) mixed with Distichlis spicata (spikegrass) and occasional patches of the shrub
Ivafrutescens {mMsh elder) and various forbs. Beyond the S. patens zone and at normal
high tide, Juncuspierardi (blackgrass) forms pure stands. At the upper edge of a marsh
inundated only by spring tides, two groups of species are common, depending on the
local rainfall and temperature. Where rainfall exceeds evapotranspiration, salt-tolerant
species give way to less tolerant species, such as Panicum vir^atum (switchgrass),
Phrapimites australis (common reed), Limonium caroUnianum (sea lavender). Aster
spp. (asters), and Tripflochin maritima (arrow grass). On the southeastern New Eng-
land coast where evapotranspiration may exceed rainfall during the summer, salts can
accumulate in these upper marshes, and salt-tolerant halophytes such as Salicornia spp.
(saltwort) and Batis maritima flourish. Bare areas with salt efflorescence are common.
Other features of New England salt marshes include well-flushed mosquito ditches
lined with tall S. alterniflora and salt pannes containing short-form S. alterniflora.

Crain et al. (2004) used greenhouse and field transplants along a New England
coast to compare biotic and abiotic factors influencing salt marsh plants. Salt marsh

Tidal Salt Marshes 271

Common Reed

. . Marsh eWef ’ Switchorass ’ 'X

' Stunted cord-grass Panne

Blackgrass

HI. I .

Salt meadow
cord-grass

HIGH MARSH

LOW

MARSH

WTtRTIDAL

ZONE

Jiik. U

Figure 8.6 Idealized zonatien of communities on a typicai North Atlantic salt marsh. The
location of the different plant associations is strongiy infiuenced by smail differences in
elevatien abeve the mean high water ievel. (After Dreyer and Niering, 1995)

plants transplanted into freshwater marsh conditions without competitors grew better
than in the salt marsh. However, when salt marsh plants were transplanted into a
freshwater marsh with neighboring plants, they were outcompeted by the freshwater
marsh plants. The authors surmised that plants in environmental extremes, such as salt
marshes, are determined by their tolerance to physiological stress while inland wetland
plant occurrence is dictated by the plant’s competitive ability.

Characteristic patterns of vegetation found in other salt marshes are shown in
Figure 8.7. South of the Chesapeake Bay along the Atlantic Coast, salt marshes typical
of the Coastal Plains appear (Fig. 8.7a). These marshes are similar in zonation to those
in New England except that (1) tall S. alterniflora often forms only in very narrow
bands along creeks, (2) the short form of S. alterniflora occurs more commonly in
the wide middle zone, and (3) Juncus roemerianus (black rush) replaces J. £ferardi
in the high marsh. At maturity, low and high marsh areas are approximately equal.
The low marsh is almost entirely S. alterniflora, tall on the creek bank and shorter
behind the natural levee as elevation gradually increases in an inland direction. It may
contain small vegetated or unvegetated ponds and mud barrens. The high marsh is
much more diverse, containing short S. alterniflora intermixed with associations of
Distichlis spicata, Juncus roemerianus, and Salicornia spp .

a. Southeast Atlantic coast salt marsh

tidal

creek

low marsh

high marsh

upland

b. Eastern Gulf of Mexico coast salt marsh

Figure 8.7 Zonation of vegetation in typical sait marshes: (a) southeastern U.S. Atlantic
Coast; (b) eastern and northern Gulf of Mexico; (c) northern France. MHT - mean high
tide; MSL — mean sea levei. ((a) After Wiegert and Freeman, 1990; (b) after Montague and
Wiegert, 1990; (c) after LeFeuvre and Dame, 1994)

272

Tidal Salt Marshes 273

CL European salt marsh

coastline low marsh miadie marsh

marsh

Figure 8.7 {Continued)

Along the Mississippi and northwest Florida coasts, J. roemerianus is found in
extensive monocultures (Fig. 8.7b). There is often a fringe of S. alterniflom along the
seaward margin, followed, in an inland direction, by large areas of tall and short J. roe-
merianus. Mixtures of S. patens and D. spicata line the marsh on the landward edge,
and Salicornia spp. can be found in small areas such as berms where salt accumulates.
Along the northern Gulf Coast, S. patens is the dominant species, occurring in a broad
zone inland of the more salt-tolerant S. alterniflora. More than 200,000 ha of coastal
marsh in Louisiana are dominated by S. patens.

In Europe, a totally different salt marsh is found, at least compared to the eastern
United States marshes (Fig. 8.7c). One of the most notable features is that the inter-
tidal zone between high tide and mean high tide is sparsely covered if it is vegetated at
all in Europe, whereas it is dominated by S. alterniflora in the United States (Lefeu-
vre and Dame, 1994). So much of what would be called the low marsh in Europe is,
in fact, a mud flat or sparsely vegetated. In Europe, the salt marshes that are studied
are mostly between mean high tide and spring tide. The cordgrass found in Europe
is generally S. an^lica or S. townsendiii, and it is found in a relatively narrow band.
In the last few decades, the clonal grass Elymus athericus has spread into the middle
and low marshes of many European salt marshes. This invasive plant, because of its
large size, traps macrodetritus on the marsh, limiting its export to the adjacent estuary
(Bouchard and Lefeuvre, 2000; Lefeuvre et ah, 2003; Valery et al., 2004) and has a
significant impact on salt marsh biodiversity (Petillon et ah, 2005).

Salt marshes closer to the polar regions are less well understood. Funk et al. (2004)
found that elevation, conductivity, and soil ion composition all contributed to plant

274 Chapter 8 Tidal Marshes

cover and species composition in an Alaskan salt marsh. As elevation increased, salinity
decreased, resulting in increased plant species richness. At the lowest marsh eleva-
tions, only PuccineUa phry£ianodes occurred. Midelevation sites were dominated by
Carex subspathaceae, and high-elevations sites had the highest cover with as many
as 16 species, including Dupontia fischeri and Eriophorum an^ustifolium. Zhu et al.
(2008) reported that biomass in coastal tundra marshes in eastern Antarctica were
dominated by a combination of algae, moss, cyanobacteria, and bacteria.

Consumers

Salt marshes, whose features are characteristic of both terrestrial (aerobic) and aquatic
(anoxic) environments, provide a harsh environment for consumers. Salt is an addi-
tional stress with which they must contend. In addition, the variability of the environ-
ment through time is extreme . The dominant plant food source for marsh consumers
is generally a marsh grass, which is usually limited in its nutritional value. Considering
all these limitations, the number of consumers in the salt marsh is surprisingly diverse.

Many faunal species, particularly vertebrate taxa, utilize tidal marshes (both
salt and freshwater) as a component of a larger set of coastal ecosystems. There
are, however, many reported species that are considered endemic to tidal marshes.
In a global review of terrestrial vertebrates and their occurrence in tidal marshes,
Greenberg et al. (2006) found 25 species (or subspecies) that were endemic to tidal
marshes. Interestingly, nearly all of the species were restricted to North America.
Sampling bias may explain some of this; however, very few records of tidal endemics
were reported in well-studied regions, such as Europe and Australia. Another possible
factor is the large area of tidal marshes present in North America, and the higher
occurrence of endemism is a reflection of a species-area relationship. While these
factors or others may contribute to the high North American endemism, currently
there is no comprehensive theory for this phenomenon.

It is convenient to classify consumers according to the type of marsh habitat they
occupy, although the animals, especially in the higher trophic levels, move from one
habitat to another. The marsh can be divided into three major habitats: an aerial
habitat — the above-ground portion of the macrophytes, which is seldom flooded; a
benthic habitat — the marsh surface and lower portions of the living plants; and an
aquatic habitat — the marsh pools and creeks (Fig. 8.8).

Aerial Habitat

The aerial habitat is similar to a terrestrial environment and is dominated by insects and
spiders that live in and on the plant leaves. This is the grazing portion of the salt marsh
food web. The most common leaf-chewing organisms in salt marshes in the eastern
United States are the arthropod Orchelimum, the weevil Lissorhoptrus, and the square-
back crab Sesarma. In addition, there are abundant sap-sucking insects {Prokelisia
mar£[inata, Delphacodes detecta) that ingest material translocated through the plant’s
vascular tissue or empty the contents of mesophyll cells. Numerous carnivorous insects
are also found in this habitat. Pfeiffer and Wiegert (1981) listed 81 species of spiders
and insects in North Carolina, South Carolina, and Georgia Spartina marshes.

Tidal Salt Marshes 275

Aerial Habitat

exchange
■4 — with
estuary

[Respiratory
I energy
— loss

Aquatic H^itat

Figure 8.8 Salt marsh food web, showing the major producer and consumer groups of the
aerial habitat, benthic habitat, and aquatic habitat. (After Montague and Wiegert, 1990)

Salt marshes support large populations of wading birds, including egrets, herons,
willets, and even wood stork and roseate spoonbills. Coastal marshes also support
vast populations of migratory waterfowl, including the mallard {Anas platyrhynchos),
American wigeon {Anas americana), gadwall {Anasstrepena), redheads {Aythya amer-
icana)^ and teals {Anas discors and A. crecca). Black duck {Anas ruhripes) is a perma-
nent resident in many marshes, as are a number of songbirds.

A number of birds, including the marsh wren ( Cistothorus palustris) and the
seaside sparrow {Ammodmmus maritimus), laughing gulls {Larus atricilla), and
Forster’s {Sterna for steri) and common terns (5. hirundo), feed and nest in the marsh
grasses. Wrens feed primarily on insects, and the sparrows apparently feed on the
marsh surface, eating worms, shrimp, small crabs, grasshoppers, flies, and spiders.
The clapper rail {Rallus longirostris) is another permanent marsh resident, feeding
primarily on cutworm moths and small crabs. Many nonpermanent insectivorous
birds forage in the salt marsh periodically, entering from adjacent fresher marshes,
beaches, and upland habitats or migrating through. These include the sharptailed
sparrow {Ammodramus caudacutus), swallows {Tachycineta bicolor, Hirundo rustica,
and Stelgidopteryx serripennis), red-winged blackbirds {Agelaius phoeniceus), and
various gulls.

Migratory waterfowl use coastal marshes extensively, mostly as wintering grounds,
but also as stopover areas during fall and spring migrations. In some areas, geese
or duck flocks numbering in the hundreds of thousands denude coastal marshes.
Repeated and intense herbivory, especially when followed by high water levels, salinity
extremes, or extended drought, may result in the formation of mud flats or shallow
open-water ponds.

276 Chapter 8 Tidal Marshes

Benthic Habitat

Probably less than 10 percent of the above-ground primary production of the salt
marsh is grazed by aerial consumers. Most plant biomass dies and decays on the marsh
surface, and its energy is processed through the detrital pathway. The primary con-
sumers are microbial fungi and bacteria. These organisms, in turn, are preyed on by
meiofauna in the decaying grass, the surface microfilm of the marsh, and the decaying
bases of plant shoots. Most of these microscopic organisms are protozoa, nematodes,
harpacticoid copepods, annelids, rotifers, and larval stages of larger invertebrates. The
larger invertebrates on the marsh surface are of two groups, foragers (deposit feeders)
and filter feeders. In a general sense, they are considered aquatic because most have
some kind of organ to filter oxygen out of water. Foragers include polychaetes, gas-
tropod mollusks such as Littorina irromta and Mdampus bidentatus^ and crustaceans
such as Uca spp., the blue crab {Callinectes sapidus), and amphipods. These organisms
browse on the sediment surface, ingesting algae, detritus, and meiofauna. The filter
feeders, such as the ribbed mussel {Geukensia demissus) and the oyster {Cmssostrea
virjjinica), filter particles out of the water column.

Aquatic Habitat

Animals classified as aquatic overlap with those in the benthic habitat. For convenience,
we include in this group animals in higher trophic levels (mostly vertebrates) and
migratory organisms that are not permanent residents of the marsh. Few fish species
are permanent residents of the marsh. Most feed along the marsh edges and in small,
shallow marsh ponds and move up into the marsh on high tides. Werme (1981) found
30 percent of silverside {Menidia extensa) and mummichog {Fundulm heterocUtus) in
a North Atlantic estuary up in the marsh at high tide. Fish common in small salt marsh
ponds in Louisiana include sheepshead minnow {Cyprinodon varie^atus varie^atus)^
diamond killifish {Adinia xenica)^ tidewater silverside {Mmidia beryUina)^ Gulf killi-
fish (Fundulus grandis)^ and sailfin molly [Poecilia latipinna). Shrimp {Pmaeus spp .)
and blue crabs ( Callinectes sapidus) are also common. Most other species use the marsh
intermittently for shelter and for food but range widely. Many fish and shellfish spawn
offshore or upstream and, as juveniles, migrate into the salt marsh, which offers an
abundant food supply and shelter. As subadults, they migrate back into the estuary
or offshore. This group of migratory organisms includes more than 90 percent of the
commercially important fish and shellfish of the southeastern Atlantic and Gulf coasts.

Mammals

Two mammals in North American salt marshes deserve attention because of their
impact on the marshes. The muskrat ( Ondatra zibethicus) is native to North America;
the coypu or nutria {Myocastor coypus) is an exotic species introduced from South
America. Both prefer fresh marshes but are also found in salt marshes. In Louisiana,
the muskrat appears to have been displaced by the nutria from its preferred freshwater
habitat into saline marshes. Both mammals are voracious herbivores that consume
plant leaves and shoots during the growing season and dig up tubers during the winter.
They destroy far more vegetation than they ingest and are responsible for eat-outs

Tidal Salt Marshes 277

that degrade large areas of marsh. These areas recover extremely slowly, especially in
the subsiding environment of the northern Gulf Coast. In European marshes, it is
common to have domestic animals (e.g., cattie, sheep, or goats) grazing in coastal
salt marshes (Bouchard et al., 2003). This grazing has a profound effect on the plant
communities and zonation that develops in these marshes.

Ecosystem Function

Major points that have been demonstrated in several studies about the functioning of
salt marsh ecosystems include the following five:

1 . Primary productivity of macrophytes is high in much of the salt

marsh — almost as high as in subsidized agriculture. This high productivity is
a result of subsidies in the form of tides, nutrient import, and abundance of
water that offset the stresses of salinity, widely fluctuating temperatures, and
alternate flooding and drying.

2. Although the biomass of edaphic algae is small, algal production sometimes
can be as high as or higher than that of the community’s macrophytes,
especially in hypersaline marshes.

3 . Direct grazing of vascular plant tissue is a minor energy flow in the salt
marsh, but grazing on edaphic and epiphytic algae is a significant source of
high-quality food energy for meio- and macro-invertebrates.

4. Fungi and bacteria are primary consumers that break down and transform
indigestible plant cellulose (detritus) into protein-rich microbial biomass for
consumers. This detrital pathway is a major flow of energy utilization in the
salt marsh.

5. Salt marshes have been shown at times to be both sources and sinks of
nutrients, particularly nitrogen.

Primary Productivity

Tidal marshes are among the most productive ecosystems in the world, annually pro-
ducing up to 80 metric tons per hectare of plant material (8,000 g-wet weight
yr^^ ) in the southern Coastal Plain of North America. The three major autotrophic
units of the salt marsh are marsh grasses, mud algae, and phytoplankton of the tidal
creeks. Extensive studies of the net primary production have been conducted in salt
marshes, especially along the Atlantic and Gulf coasts of the United States. A com-
parison of some of the measured values of net above-ground and below-ground pro-
duction is given in Table 8.3. Above-ground production varies widely, from as little as
410 g yr^^ in a Normandy salt marsh to a high of 4,200 g yr“^ in a Louisiana

Spar tina patens miLrsh. Below-ground production is difficult to measure and can be
much higher than above-ground production. Productivity of salt marshes is often
higher along creek channels and in low or intertidal marshes than in high marshes
because of the increased exposure to tidal and freshwater flow. These conditions also

278 Chapter 8 Tidal Marshes

Table 8.3 Net primary productivity estimates of salt marshes and dominant piant species

Species

Aboveground Net
Primary Production
(g m-2 yr-i)

Belowground Net

Primary Production

(g m“^ yr“^) Source

Louisiana

Distichlis spicata

1,162-1,291

White etal. (1978)

Juncus roemerianus

1,806-1,959

Spartina alterniflora

1,473-2,895

Spartina patens

1,342-1,428

Distichlis spicata

1,967

Hopkinson et al. (1980)

Juncus roemerianus

3,295

Spartina alterniflora

1,381

Spartina cynosuroides

1,134

Spartina patens

4,159

Aiabama

Juncus roemerianus

3,078

7,578 Stout (1978)

Spartina alterniflora

2,029

6,218

Mississippi

Juncus roemerianus

1,300

de la Cruz (1974)

Distichlis spicata

1,072

Spartina alterniflora

1,089

Spartina patens

1,242

Normandy (France)

Spartina anglica/

1,080 (nongrazed)

Lefeuvre et al. (2000)

Saiicornia/Suaeda

410 (grazed)

maritima low marsh

1,990 (nongrazed)

high marsh

550 (grazed)

Mediterranean Sea

Rhone River Delta

Ibanez et al. (1999)

(France)

Sarcocornia fruticosa

1123-1262

Ebro Delta (Spain)

Curco et al. (2002)

Arthocnemum

macrostachyum

190

Sarcocornia fruticosa

580

950

A. macrostachyum, S.

840

340

fruticosa

produce the taller forms of Spartina, as discussed earlier. Below-ground production
is sizable — often greater than aerial production (Table 8.3). Under unfavorable soil
conditions, plants seem to put more of their energy into root production. Hence,
rootsishoot ratios seem to be generally higher inland than at streamside locations.

The productivity of edaphic algae was summarized by Sullivan and Currin (2000).
Annual benthic algal production, as measured in a number of studies, ranges ftom 28 g
C yr“^ in a Gulf of Mexico coast Juncus roemerianus marsh to 341 g C yr“^
in a southern California Jaumm carnosa marsh (Table 8.4). Benthic algal produc-
tion increases in a southerly direction along the Atlantic coast, but is lowest on the

Tidal Salt Marshes 279

Table 8.4 Comparison of annual benthic microalgal production (g C yi^^) and ratio
of annual benthic microalgal to vascular plant net aerial production (BMP/VPP) in
different salt marshes of the United States

State

Algal productivity,
g C ypl

BMP/VPF? %

Reference

Massachusetts

105

Van Raalte (1976)

Delaware

61-99

Gallagher and Daiber (1974)

South Carolina

98-234

12-58

Pinckney and Zingmark (1993)

Georgia

200

Pomeroy (1959)

Georgia

150

Pomeroy et al. (1981)

Mississippi

28-151

10-61

Sullivan and Moncreiff (1988)

Texas

8-13

Hall and Fisher (1985)

California

185-341

76-140

Zedler (1980)

Source: Sullivan and Currin (2000)

Gulf coast. Much of the algal production on the east and west coasts of the United
States occurs when the overstory plants are dormant. On the Atlantic and Gulf coasts,
the productivity of algae is 10 to 60 percent of vascular plant productivity. Zedler
(1980), however, found that algal net primary productivity in southern Galifornia was
76 to 140 percent of vascular plant productivity. She hypothesized that the arid and
hypersaline conditions of southern Galifornia favor algal growth over vascular plant
growth. Algae are important components of the salt marsh food web, so much so that
Kreeger and Newell (2000) stated:

We question the paradigm that salt marshes have “detritus-based food webs”

(Odum, 1980), considering that the bulk of secondary production by metazoans
could actually be linked to primary production by the microphytobenthos rather
than through either direct (herbivory) or indirect (detrivory) linkages to primary
production by vascular plants.

Variations in productivity on the local scale result from complex interactions of soil
anoxia, soluble sulfide, and salinity (Mendelssohn and Morris, 2000). Although water
appears plentiful, the concentration of dissolved salt makes the salt marsh environment
similar in many respects to a desert. The “normal” water gradient is from plant to
substrate. To overcome the osmotic influence of salt, plants must expend energy to
increase their internal osmotic concentration in order to take up water. As a result,
numerous studies confirm that plant growth is progressively inhibited by increasing
salt concentrations in the soil. This is true even for the salt-tolerant species of the salt
marsh, and the salinity effects may be subtle. For example, Morris et al. (1990) showed
that the year-to-year variation in marsh production at a single site on the East Goast
of the United States was correlated with the mean summer water level, which they
equated with soil salinity. (Soil salinity was inversely correlated with the frequency of
marsh flooding in this study. )

Another factor limiting production is the degree of anaerobiosis of the substrate.
Vascular plants, even those that have developed adaptations to anaerobic conditions.

280 Chapter 8 Tidal Marshes

grow best in aerobic soils. Many effects of anaerobiosis have been documented:
reduced energy availability as the aerobic respiratory pathway is blocked, reduced
nutrient uptake, the accumulation of toxic sulfides in the substrate, and changes
in the availability of nutrients. Salt inhibition and oxygen depletion frequendy
occur together. Spartina grows shorter in the inland marsh because its drainage is
poor; hence oxygen deficits are severe. Salt may concentrate in this environment.
The primary result of poor drainage in inland salt marshes, however, is apparendy
a dramadcally lower soil redox potential, which in turn leads to elevated sulfide
concentradons. Although S. alterniflora is able to midgate the toxic effects of sulfide
to some extent through its ability to transport oxygen through the root system to
the rhizosphere and by the enzymadc oxidation of sulfides, its growth is inhibited
when the interstitial soluble sulfide concentration exceeds 1 mM sulfide (Bradley and
Dunn, 1989; Koch et ak, 1990).

Primary producdvity also contributes to sediment accretion in salt marshes, and
there is increasing interest in the ability of salt marshes to withstand reladve rises in sea
level. Because of their locadon, salt marshes are constandy adjusdng to maintain an
equilibrium near mean sea level. Morris et al. (2002) demonstrated the importance of
primary producdvity for increased sediment accredon. By experimentally increasing
producdvity in a South Carolina salt marsh, sediment accredon in the marsh was also
enhanced. Salt marshes tend to be most producdve at elevadons just below mean
high tide (Fig. 8.9). However, in terms of long-term response to rapidly rising sea
levels, these lower marshes may be incapable of accreting sediment quickly enough to

Figure 8.9 Above-ground net primary productivity of salt marshes as a function of elevation
below mean high tide during peak growing season of June and July. Data are for high (open
circles) and low (solid circles) Spartina alterniflora salt marshes. (After Morris et al., 2002)

Tidal Salt Marshes 281

keep pace with rising water levels. Marshes at slightly higher elevations with abundant
supplies of sediment are the most likely ones to acclimate to rapidly rising sea levels
and maintain their position.

Decomposition and Consumption

Since John Teal’s seminal publication on energy flow in the salt marsh system (Teal,
1962), salt marshes have been considered detrital systems. Almost three-quarters of
the primary production in the salt marsh ecosystem is broken down by bacteria and
fimgi. In his study of energy flow within the salt marsh environment. Teal (1962)
estimated that 47 percent of the total net primary productivity was lost through respi-
ration by microbes. It was largely assumed that the rich secondary productivity of estu-
aries was fueled by a detritus food web. With the development of new techniques, such
as multiple stable isotope fractionation, these early assumptions have been questioned,
and a refined and quite different picture of decomposition and secondary production
has emerged. With a few exceptions such as salt marshes in Mediterranean-type cli-
mates, primary production is dominated by emergent spermatophytes, usually grasses.
When they senesce, the soluble organic contents are rapidly flushed from their tissues.
This labile soluble organic matter from both living and decomposing salt marsh vege-
tation (which may be as much as 25 percent of the initial dry weight of the dying grass)
is an important energy source for microorganisms in the marsh and the adjacent estu-
ary (Wilson et ah, 1986; Newell and Porter, 2000). The remaining 75 percent of dead
vegetation biomass is largely composed of refractory structural lignocellulose that is
indigestible by all but a few metazoans. Ideas about the fate of the vegetation biomass
have changed. Two key conclusions about the process of decomposition are described
next.

1 . The initial secondary producers, or decomposers, on epibenthic marsh grass
stems are ascomycetous fungi. These fungi may reach a biomass equal to 3
(summer) to 28 (winter) percent of live Spartina alterniflora standing crop.
Most of this biomass occurs in standing dead grass or on the marsh surface.

In South Atlantic coastal marshes, fungal productivity is 10 times greater in
winter than in summer. In contrast, most of the bacterial biomass is found in
the sediment surface microlayer. Productivity of bacteria is twice fungal
productivity in summer but only one-tenth as great in winter. The conversion
efficiency of grass biomass to fungal biomass can be as high as 50 percent
(Newell and Porter, 2000).

2. There appear to be at least three decomposer groups, (a) Fungi are the major
decomposers of the epibenthic standing dead grass; (b) aerobic bacteria in
the surface microlayer decompose the decayed grass leaf shoots that are
shredded by gastropods and amphipods and fall to the marsh surface; and

(c) anaerobic bacteria, a third group of decomposers in deeper anoxic
sediments, are able to use electron acceptors other than oxygen to
metabolize. Prime among these are sulfate reducers, which may oxidize a
major proportion of the underground senescent root and rhizome biomass.

282 Chapter 8 Tidal Marshes

During the decomposition process, the nitrogen content of the grass/fungal/
bacterial brew increases. This is due, in part, to the low C:N ratio of bacterial decom-
posers compared to raw grass tissue. It was assumed for many years that nitrogen
enrichment made the decaying plant material a nutritionally better food supply for
consumers; in more recent studies, however, it was determined that much of the
nitrogen is bound in refractory compounds in the decaying grass (Teal, 1986). The
nutritious bacterial population is kept at low concentrations by metazoan grazing.

These discoveries about the decomposition process in marsh macrophytes have
led to a reevaluation of the source of energy for the abundant consumer population
found in tidal marshes and their associated tidal creeks. Although much of the change
consists of elaboration and clarification of the detrital process, a major shift has been
toward a much greater role for algae, both phytoplankton and especially edaphic algae,
as major flows of energy in the salt marsh food web. Vascular plants are still the major
source of organic carbon, but few metazoans can assimilate this cellulose-rich mate-
rial. Hence, direct grazers are limited to several species of herbaceous insects, which
collectively consume less than 10 percent of plant production. The ribbed mussel
Gmkensia demissa is an exception to this generalization. It has been shown to assimi-
late aseptic detrital cellulose with an efficiency of up to 15 percent. Kreeger and Newell
(2000) suggest that the mussel must either possess endogenous cellulases or contain
a vigorous gut flora capable of cellulose breakdown.

Decay begins with fungal decomposition of the aerial parts of the senescent vas-
cular plants. Epiphytic algae growing on the lower parts of the grass culms are also a
part of this detrital brew. The complex is ingested and shredded by gastropod snails,
such as Littorina irromta and perhaps amphipods. In a microcosm experiment, the
snails had the capacity to ingest 7 percent of their weight of naturally decayed leaves
per day and assimilate it with an efficiency of about 50 percent. The epiphytic algae
are also ingested and assimilated by amphipods and other organisms grazing on the
dead leaf surfaces.

The finely shredded grass/fungal/algal material that falls to the marsh surface is
infected by aerobic bacteria that continue the process of decomposition. Also part of
this mixture is the algal community, largely diatoms, growing on the marsh surface.
This complex is consumed by benthic meiofaunal and macrofaunal deposit feeders.
Primary among the meiofauna are nematodes; also feeding on the surface are harpacti-
coid copepods, amphipods, polychaetes, turbellarians, ostracods, foraminifera, and
gastroliths. The larger consumers in this group include fiddler crabs, snails, poly-
chaetes, oligochaetes, and some bivalves.

Finally, some of the finely decomposed organic material on the surface microlayer
is periodically suspended by winds and currents, where it mixes with the phytoplankton
growing in the water. For example, as much as 25 percent of the suspended algae have
been found to be edaphic species (MacIntyre and Cullen, 1995). This sestonic mix of
bacterial/organic fragments, free-living bacteria, and algae is consumed by suspension
feeders, especially benthic suspension feeders such as bivalve mollusks and oligochaete
annelids. Also active are zooplankton, although they probably do not process as much

Tidal Salt Marshes 283

material as the benthic bivalves. The meio- and macrofauna feeding on algae, fungi,
and bacteria are in turn consumed by animals in the higher trophic levels (Fig. 8.8).

Organic Export

A central paradigm of salt marsh ecology has long been the outwellii\0 hypothesis^ which
was first enunciated by E. P. Odum in 1968. The hypothesis was based, in part, on a
salt marsh energy flow analysis presented by John Teal at the first salt marsh confer-
ence, held in 1958 at the University of Georgia Marine Laboratory on Sapelo Island,
Georgia (published as Teal, 1962). Odum (1968) described salt marshes as “primary
production pumps” that feed large areas of adjacent waters, and he compared the
flow of organic material and nutrients from salt marshes to the upwelUn^ of deep
ocean water, which supplies nutrients to some coastal waters. Teal (1962) hypoth-
esized that salt marshes exported organic material and energy primarily as detritus
from the marsh surface. In the intervening years, there have been many attempts to
measure this export. Teal’s (1962) energy flow analysis estimated that about 45 per-
cent of net primary production was exported from the salt marsh. Nixon’s (1980)
summary agreed in that most studies showed an export of dissolved and particulate
material, in an amount that could account for about 10 to 50 percent of phytoplankton
production in coastal and estuarine waters.

Childers et al. (2000) pointed out that the original hypothesis was ambiguous in
that it equated salt marsh export to coastal ocean import. In reality, the flows from
a salt marsh are into nearby tidal creeks, and fluxes to the coastal ocean depend on
the geomorphology of the estuary and the distance from the marsh to the coast.
Hence, salt marshes interact with nearby tidal creeks and the inner estuary, which,
in turn, exchange flows with the greater estuarine basin, which, finally, interacts with
the coastal ocean (Fig. 8.10). Failure to take these spatial factors into account has

Marsh water-column
exchange

A ►

Subbasin estuary Estuary ocean
exchange exchange

A *■ A ►

Greater estuary

Ocean

Figure 8.10 Hierarchy of estuarine-coastal landscape that includes estuarine subbasins
nested within the greater estuary, and vegetated wetland ecosystems nested within beth.
SAV -- submerged aquatic vegetation. (After Childers et al., 2000)

284 Chapter 8 Tidal Marshes

made it difficult to compare studies and is one reason for the lack of agreement in
study results.

The evidence for outwelling rests on more than organic flux data, such as
reported in the studies summarized by Nixon (1980) and Childers (1994). Hopkin-
son (1985) reported that water column respiration offshore of the Georgia barrier
islands exceeded in situ production; that is, the zone was heterotrophic, implying
that organic matter was being imported from the inshore estuaries and marshes.
Turner et al. (1979) reported that offshore, within 10 km of the coastal estuaries,
primary productivity measurements were often 10 times greater than that farther
offshore. They attributed this high productivity to outwelling of nutrients from the
estuaries.

Other evidence of outwelling comes from fishery studies. Turner (1977) found a
close correlation worldwide between commercial yield of shrimp (which are harvested
both in the estuary and offshore) and the area of estuarine intertidal vegetation. Teal
and Howes (2000) analyzed fish catch statistics dating back to 1880 from the Long
Island Sound, New York, and determined that fish catch was closely related to marsh
edge length. Since edge length is an index of accessibility to the marsh, the result
implicated salt marsh production in commercial fishery catch.

Several general factors affect the outwelling hypothesis. First, material and energy
usually flow from concentrated hot spots to lower concentration areas. Salt marshes are
hot spots of production, so it is logical to expect an outwelling of production and food
energy (E. P. Odum, 2000). Second, outwelling can be expected to be modified by
the geomorphology of the estuary and the location of a salt marsh in the estuary. Thus,
open estuaries with salt marshes close to the coast are expected to export more material
than estuaries with small coastal passes and distant marshes. Finally, salt marshes and
coastal estuaries are pulsing systems, with daily tidal variation, seasonal variations in
rainfall and river flow, and periodic severe storms. Extreme events often lead to import
or export that overwhelms the normal daily fluxes.

Salt Marsh Die-off

For the first several years at the turn of the century (2000-2005), Spartina salt
marshes in the southeastern and Gulf coasts of the United States were experiencing
major die-off, totaling more that 100,000 ha and affecting 1,500 km of coastline.
One theory presented to explain this die-off was described by Silliman et al. (2005)
as having the following sequence: a protracted and intense drought that occurred
for three to four years (bottom-up effect) followed by snails {Littomria irromta)
concentrating on the die-off borders to prolong the effect (top-down effect). In
addition, declines in blue crab populations, a major predator of the snails, of 40 to
85 percent provided synergy for the snail grazing. Essentially Silliman et al. (2005)
suggested that “drought-induced soil stress can amplify top-down control by grazers
and initiate marsh plant die-off. . . . These disturbances then stimulate the formation
of consumer fronts, leading to waves of salt marsh destruction resulting from runaway
consumption.” Such epidemic ecosystem die-offs that combine bottom-up and
top-down stresses on coastal ecosystems in a synergistic way are another example of

Tidal Freshwater Wetlands 285

an undesirable positive feedback that could occur on coastal ecosystems with any
significant climate change.

Tidal Freshwater Wetlands

Tidal freshwater wetlands are interesting because they receive the same “tidal subsidy”
as mangroves and salt marshes but without the salt stress. One would expect, there-
fore, that these ecosystems might be very productive and also more diverse than their
saltwater counterparts. As tides attenuate upstream, the wetlands assume more of the
characteristics of inland freshwater wetlands (see Chapter 10: Freshwater Marshes).
The distinction between freshwater tidal and inland wetlands is not clear-cut because
on the coast they form a continuum (Fig. 8.1). Inland from the tidal salt marshes but
still close enough to the coast to experience tidal effects, tidal freshwater marshes are
dominated by a variety of grasses and by annual and perennial broad-leaved aquatic
plants. In the United States, they are found primarily along the Middle and South
Atlantic coasts and along the coasts of Louisiana and Texas. Tidal freshwater swamps
tend to be most abundant along the farthest tidal extent of coastal rivers, particu-
larly those rivers with low gradients and high discharge. Most of the extensive tidal
freshwater forests in the United States occur along the southeastern coastline (Mary-
land to Texas). Estimates of tidal freshwater wetlands in the United States range from
400,000 ha along the Atlantic Coast to 819,000 ha for the conterminous United
States (Table 8.1). The extent of tidal freshwater swamps in the United States is less
certain, but 200,000 ha has been is estimated for the southeast U.S. coastline (Field
et ah, 1991). The uncertainty in the estimates is related to where the line is drawn
between tidal and nontidal areas. Tidal freshwater marshes can be described as interme-
diate on the continuum from coastal salt marshes to freshwater marshes. Because they
are tidally influenced but lack the salinity stress of salt marshes, often tidal freshwater
marshes have been reported to be very productive ecosystems, although a considerable
range in their productivity has been measured. Elevation differences across a freshwa-
ter tidal marsh correspond to different plant associations. These associations are not
discrete enough to call communities, and the species involved change with latitude.
Nevertheless, they are characteristic enough to allow some generalizations.

Vegetation

Marsh Vegetation

On the Atlantic Coast of the United States (Fig. 8.11a, b), submerged vascular
plants, such as Nuphar advena (spatterdock), Elodea spp. (waterweed), Potamojjeton
spp. (pondweed), and Myriophyllum spp. (water milfoil), grow in the streams and
permanent ponds. The creek banks are scoured clean of vegetation each fall by the
strong tidal currents, and they are dominated during the summer by annuals, such as
Polypionum punctatum (water smartweed), Amaranthus cannabinus (water hemp),
and Bidens laevis (bur marigold). The natural stream levee is often dominated by
Ambrosia trifida (giant ragweed). Behmd this levee, the low marsh is populated with

Open Water Low Marsh High Marsh Wooded Swamp Forested Upland

-Sedges Rushes —
Martina sp.

-^Fbntedwa cordata
Nupharaitvena
Rooted aquatics

j^ha

■Hbiacusmoacheutos
— Impatiens c^peoss

\-R>lygonum anfolium —
-i-Fb1ygonumsp.

- f^tandra virginica

Channel

Stream Bank/
Levee

High Marsh

Pond-Like

Ponds

Figure 8.11 Cross sections across typical freshwater tidal marshes, showing elevation
changes and typicai vegetation: (a) and (b) Atlantic Coast marshes; (c) new marsh in the
Atchafalaya Deita, Louisiana, ((a) After W. E. Odum et al., 1984; (b) after Simpson et al.,
1983; (c) after Gosseiink et ai., 1998)

286

Tidal Freshwater Wetlands 287

Channel Levee Typha Marsh Interior Flats

Mean low

Salix nigra

■Beochansparvula
; C^perusdifformis

"^ha latiMia

Figure 8.11 {Continued)

broad-leaved monocotyledons, such as Peltandra vir^inica (arrow arum), Pontederia
cordata (pickerelweed), and Sagittaria spp. (arrowhead).

Typically, the high marsh has a diverse population of annuals and perennials.
W. E. Odum et al. (1984) called this the “mixed aquatic community type” in the
Mid-Adantic region. Leek and Graveline (1979) described a “mixed annual” associa-
tion in New Jersey while Caldwell and Crow ( 1992 ) described the vegetation of a tidal
freshwater marsh in Massachusetts. Generally, the areas were dominated early in the
season by perennials, such as arrow arum. A diverse group of annuals — Bidms laevis,
Polygonum arifoUum (tear-thumb) and other smartweeds. Pika pumila (clearweed).
Hibiscus coccimus {rose mallow), Acnida cannabina^ and others — assumed dominance
later in the season. In addition to these associations, there are often almost pure stands
of Zizania aquatica (wild rice), Typha spp. (cattail), Zizaniopsis miliacea (giant cut-
grass), and Spartina cynosuroides (big cordgrass). In the northern Gulf of Mexico,
arrowheads {Sa^ittaria spp.) replace arrow arum {Peltandra spp.) and pickerelweed
{Pontedaria cordata) at lower elevations. Visser et al. (1998) described three vegeta-
tion associations in this area:

1. Bulltongue {Sajjittaria lancifoUa) occurs with co-dominants maidencane
{Panicum hemitomon) and spikerush {Eleocharis spp.). Commonly the ferns
Thelypteris palustris 2ind Osmunda reptalis^ wax myrtle {Myrica cerifera)., and
pennywort {Hydrocotyi spp.) are also present. Fifty-two different species
occur in this association.

2. A maidencane-dominated association is widespread across the delta and
includes 55 species.

288 Chapter 8 Tidal Marshes

3. Cutgrass {Zizaniopsis miliacea) occurring with co-dominant maidencane is
relatively uncommon. It includes 20 other species.

Interestingly, rising sea level and/or surface subsidence on both the Gulf Coast
and the Atlantic Coast has resulted in vegetation shifts. Although the previously
dominant species are still present, in a Chesapeake Bay tidal freshwater wetland, for
example, the oligohahne species Spartina cynosuroides^ which was not among the
dominant species in 1974, is now second in peak biomass and fourth in importance
value (Perry and Hershner, 1999). Similarly, Visser et al. (1999) reported that the
maidencane association has decreased from 5 1 percent coverage of the tidal wedands
of Terrebonne Bay (in the Mississippi River delta) in 1968 to only 14 percent
in 1992. It has been replaced by Ekocharis baldwinii-dominiLtcd marshes, which
were uncommon in 1968 (3 percent coverage) but in 1992 covered 42 percent of
the area.

Floating Marshes

Floating marshes in the tidal reaches of the northern Gulf of Mexico are similar to the
nontidal riverine and lacustrine marshes found extensively around the globe. Large
expanses of floating marshes {Phra^mites cowwMwri marshes) have been found in the
Danube Delta for at least a century (Pallis, 1915), along the lower reaches of the Sud in
Africa (papyrus swamps; Beadle, 1974), in South America (floating meadows in lakes
of the varzea\ Junk, 1970), and in Tasmania (floating islands in the Lagoon of Islands;
Tyler, 1976). Floating marshes have also been reported in Germany, the Netherlands
(Verhoeven, 1986), England (Wheeler, 1980), and North Dakota and Arkansas in the
United States (Eisenlohr, 1972; Huffman and Lonard, 1983). In Louisiana, floating
marshes are usually floristically diverse, but different stands are dominated by Pan-
icum hemitomon with ferns and vines such as Vipfna luteola and Ipomoea sajjittata;
Sajjittaria lancifoUa with Eleocharis spp., Panicum dichotomiflorum, Bacopa monnieri,
and Spartina patens-, and Eleocharis baldwinii and Eleocharis parvula with Ludwi^ia
leptocarpa. Phyla nodiflora, and Bidens laevis (Sasser et ah, 1996). The marsh sub-
strate is composed of a thick organic mat, entwined with living roots, that rises and
falls (all year or seasonally) with the ambient water level (Swarzenski et ah, 1991).
This type of marsh is interesting in a successional sense because it appears to be an
endpoint in development; it is freed from normal hydrologic fluctuation and mineral
sediment deposition. Hence, in the absence of salinity intrusions, it appears to support
a remarkably stable community (Sasser et ah, 1995).

New Marshes

The active deltas of the Mississippi and Atchafalaya rivers are the sites of the largest
newly progading coastal deltas in the continental United States. Fresh tidal marshes
(Fig. 8 . 1 Ic) that have formed in the last 25 years on emergent islands are dominated on
the natural levees by black willow {Salix nigra'), and the extensive back-island mud flats
are dominated by common three-square sedge (Scirpus deltarum) or by arrowhead
{Sagittaria latifolia), with areas of cattail {Typha latifolia) and a seasonally variable
annual/perennial mix in between.

Tidal Freshwater Wetlands 289

Swamp Vegetation

The richness of canopy trees tends to be low in tidal freshwater forests, such as those
found in southeastern United States. At the lowest elevations, canopy species tend
to be dominated by those that can withstand long periods of inundation, such as
cypress {Taxodium distichum)^ water tupelo {Nyssa aquatica), and swamp tupelo
{Nyssa biflora). (See Chapter 11: “Freshwater Swamps and Riparian Ecosystems,”
for more descriptions of these forested wetlands.) At slightly higher elevations,
other species can become dominant, including ash {Fraxinus s'p'p.)., red maple {Acer
rubrum), sweetgum {Liquidambar styraciflua)^ American hornbeam {Carpinus
caroUniana) and sweetbay {Magnolia virginiana). Tidal freshwater swamps are
often characterized by a distinctive hummock and hollow topography in which most
of the trees are limited to the hummocks and hollows are sparsely vegetated to
unvegetated by trees. Because these forests can have fairly open canopies, subcanopy
and understory vegetation can be extensive and species rich. In tidal swamps along the
Pamunkey River in Virginia, Rheinhardt (1992), found spicebush {Lindera benzoin).,
winterberry {Ilex verticillata), C. caroUniana, and Ilex opaca to be among the
most dominant subcanopy species. Understory dominants included halberd -leaved
tearthumb {Folypfonum arifolium), lizard’s tail {Saururus cernuus), and sedges {Carex
spp.). Along the forested tidal reaches of the Suwannee River on the Gulf Coast
of Florida, dominant subcanopy vegetation was pumpkin ash {Fraxinus profunda),
Carolina ash {Fraxinus caroUniana), and wax myrtle {Morelia cerifera). Common
understory vegetation included variable panic grass {Dicanthelium commutatum) ,
string lily {Crinum americanum), S. cernuus, and Carex spp.

Seed Banks

The species composition of a tidal freshwater wetland does not appear to depend on
the availability of seed in particular locations. Seeds of most species are found in almost
all habitats, although the most abundant seed reserves are generally from species
found in that vegetation zone (Whigham and Simpson, 1975; Leek and Simpson,
1987, 1995; Baldwin et ah, 1996). They differ, however, in their ability to germi-
nate under the local field conditions and in seedling survival. Flooding is one of
the main controlling physical factors. Many of the common plant species seem to
germinate well even when submerged — for example, Peltandra vir^inica and Typha
latifolia — whereas others — such as Impatiens capensis, Cuscuta ^ronovii, and Poly-
gonum arifolium — show reduced germination. Baldwin et al. (2001 ) manipulated the
hydrology in a series of experiments examining freshwater tidal marshes along the
Patuxent River in Maryland and found that increasing flood depths by 3 to 10 cm can
significantly reduce species richness and plant growth. In particular, shallow flooding
early in the growing season reduced the germination of annuals.

Competitive factors also play a role in vegetation assemblages. Arrow arum and
cattail, for example, produce chemicals that inhibit the germination of seed; and shad-
ing by existing plants is apparently responsible for the inability of arrow arum plants
to become established anywhere except along the marsh fringes. Some species {Impa-
tiens capensis, Bidens laevis, and Polygonum arifolium) are restricted to the high marsh

290 Chapter 8 Tidal Marshes

because the seedlings are not tolerant of extended flooding. Seed bank strategies differ
in different zones of the marsh. The seeds of most of the annuals in the high marsh
germinate each spring so that there is little carryover in the soil. In contrast, perennials
tend to maintain seed reserves. The seeds of most species, however, appear to remain
in the soil for a restricted period. In one study, 31 to 56 percent of the seeds were
present only in surface samples, and 29 to 52 percent germinated only in sediment
samples taken in early spring (Leek and Simpson, 1987). The complex interaction of
all these factors has not been elucidated to the extent that it is possible to predict what
species will be established where on the marsh.

In addition to vascular plants, phytoplankton and epibenthic algae abound in
freshwater tidal marshes, but relatively little is known about them. In one study
of Potomac River marshes, diatoms (bacillariophytes) were the most common
phytoplankton, with green algae (chlorophytes) comprising about one-third of the
population and blue-green algae (cyanobacteria) present in moderate numbers.
The same three taxa accounted for most of the epibenthic algae. Indeed, many
of the algae in the water column are probably entrained by tidal currents off the
bottom. In a study of New Jersey tidal freshwater marsh soil algae, Whigham et al.
(1980) identified 84 species exclusive of diatoms. Growth was better on soil that was
relatively mineral and coarse, compared with growth on fine organic soils. Shading
by emergent plants reduced algal populations in the summer months. In nontidal
freshwater marshes, algae epiphytic on emergent plants and litter made important
contributions to invertebrate consumers (Campeau et al., 1994). Algal biomass is
probably two to three orders of magnitude less than peak biomass of the vascular
plants, but the turnover rate is much more rapid.

Consumers

Coastal freshwater wetlands are used heavily by wildlife. The consumer food chain is
predominantly detrital, and benthic invertebrates are an important link in the food
web. Bacteria and protozoa decompose litter, gaining nourishment from the organic
material. It appears unlikely that these microorganisms concentrate in large enough
numbers to provide adequate food for macroinvertebrates. Meiobenthic organisms,
primarily nematodes, comprise most of the living biomass of anaerobic sediments.
They probably crop the bacteria as they grow, packaging them in bite-sized portions
for slighdy larger macrobenthic deposit feeders. In coastal freshwater marshes,
the microbenthos is composed primarily of amoebae (thecamoebinids, a group of
amoebae with theca casings). This is in sharp contrast to more saline marshes in
which foraminifera predominate. The slightly larger macrobenthos is composed of
amphipods, especially Gammarus fasciatus^ oligochaete worms, freshwater snails,
and insect larvae. Copepods and cladocerans are abundant in the tidal creeks. The
Asiatic clam {Corbicula fluminaea) ^ a species introduced into the United States in the
twentieth century, has spread throughout the coastal marshes of the southern states.
Caridean shrimp, particularly Palaemonetes pugio^ are common, as are freshwater
shrimp, Macrobrachium spp. The density and diversity of these benthic organisms

Tidal Freshwater Wetlands 291

are reported to be low compared with those in nontidal freshwater wetlands, perhaps
because of the lack of diverse bottom types in the tidal reaches of the estuary.

Where coastal forests transition from tidal to nontidal, shifts in invertebrate species
have been shown to correspond with hydrological and species shifts. At the tidal tran-
sition of the Suwannee River, Wharton et al. (1982) reported that faunal associations
changed from a brackish water snail-fiddler crab community ( Neretina-Uca) to a fresh-
water snail -crayfish ( Vivipam-Cabarus) community. Salinity and not vegetation was
the primary factor dictating infauna and epifauna taxa along the lower reaches of the
Cape Fear River in North Carolina (Hackney et ah, 2007). Common faunal groups in
tidal swamps include oligochaetes (especially Tubifrcidae and Lumbriculidae), fiddler
crabs ( Uca spp.) and grass shrimp {Palaemontes pugio) .

Nekton

Coastal freshwater wedands are important habitats for many nektonic species that use
the area for spawning, year-round food and shelter, and a nursery zone and juvenile
habitat. Fish of coastal freshwater marshes can be classified into five groups (Fig. 8.12).
Most of them are freshwater species that spawn and complete their lives within fresh-
water areas. The three main families of these fish are cyp rinds (minnows, shiners, carp),
centrarchids (sunfrsh, crappies, bass), and ictalurids (catfish). Juveniles of all species are
most abundant in the shallows, often using submerged marsh vegetation for protection
from predators. Predator species — the bluegill {Lepomis macrochirus), largemouth bass

Freshwater spedes
bluegill

largemouth bass
sunfish
warmouth
black perch

Marshes

speaes
menhaden
spot
croaker
silver perch
spotted seatrout
black dmm
summer flounder
snook
tarpon

brown shrimp

Catadromous spedes
'American eel

Semi-anadromous
& anadromous spedes
striped bass
herring
shad
sturgeon

Figure 8.12 Fish and shellfish that use tidal freshwater marshes and other coastal systems
can be classified into five groups: freshwater, estuarine, anadromous, catadromous, and
estuarine-marine.

292 Chapter 8 Tidal Marshes

{Micropterussalmoides), sunfish (Lf^owwspp.), warmouth {Lcpomisgulosus)^ and black
crappie [Pomoxis nipfromaculatus) — are all important for sport fishing. Gar {Lepisosteus
spp.), pickerel {Esox spp.), and bowfin {Amia calva) are other common predators often
found in both coastal marshes and tidal freshwater creeks.

Some oligohaline or estuarine fish and shellfish that complete their entire life cycle
in the estuary extend their range to include the freshwater marshes. Killifish {Fun-
dulus spp.) ^ particularly the banded killifish (F. diaphanus) and the mummichog {F.
heteroclitus), are abundant in schools in shallow freshwater marshes, where they feed
opportunistically on any available food. The bay anchovy {Anchoa mitchilU) and tide-
water silverside {Mmidm berylUna) are also often abundant in freshwater areas. The
latter breed in this habitat more than in saltwater areas. Juvenile hog chokers ( Trinectes
anadensi) and naked gobies {Gobiosoma bosci) use tidal freshwater areas as nursery
grounds (W. E. Odum et ah, 1984).

Anadromous species of fish, which live as adults in the ocean, or semianadromous
species., whose adults remain in the lower estuaries, pass through coastal freshwater
marshes on their spawning runs to freshwater streams. For many of these species,
the tidal freshwater areas are major nursery grounds for juveniles. Along the Atlantic
Coast, herrings {Alosa spp.) and shads {Dorosonm spp.) fit into this category. The
young of all of these species, except the hickory shad (A. mediocris), are found in peak
abundance in tidal fresh waters, where they feed on small invertebrates and, in turn,
are an important forage fish for striped bass (Morone sa-xatilis), white perch (Morone
anadensi), catfish (Ictalurus spp.), and others (W. E. Odum et ah, 1984). As they
mature late in the year, they migrate downstream to saline waters and offshore. Two
species of sturgeon {Acipenser brevirostrum and A. oxyrhynchus) were formerly impor-
tant commercially in East Coast estuaries but were seriously overfished and currendy
are rare. Both species spawn in nontidal and tidal fresh waters, and juveniles may spend
several years there before migrating to the ocean.

The striped bass is perhaps the most familiar semianadromous fish of the
Mid -Atlantic Coast because of its importance in both commercial and sport fisheries.
Approximately 90 percent of the striped bass on the East Coast are spawned in
tributaries of the Chesapeake Bay system. They spawn in spring in tidal fresh and
oligohaline waters; juveniles remain in this habitat along marsh edges, moving
gradually downstream to the lower estuary and nearshore zone as they mature.
Because the critical period for survival of the young is the larval stage, conditions in
the tidal fresh marsh area where these larvae congregate are important determinants
of the strength of the year class.

The only catadromous species in Adantic Coast estuaries is the American eel
{Anpiuilla rostrata). It spends most of its life in fresh or brackish water, returning to
the ocean to spawn in the region of the Sargasso Sea. Eels are common in tidal and
nonddal coastal freshwater areas, in marsh creeks, and even in marshes.

The juveniles of a few species of fish that are marine spawners move into freshwa-
ter marshes, but most remain in the oligohaline reaches of the estuary. Species whose
range extends into ddal freshwater marshes are menhaden {Brevoortia tyrannus), spot

Tidal Freshwater Wetlands 293

{Leiostomus xanthurus)^ croaker (Micropojjonias undulatus), silver perch {Bairdiella
chrysoura), spotted seatrout {Cynoscion nebulosus), black drum {Pojjonias cromis),
summer flounder {Paralichthys dentatus)^ snook {Cmtropomus undecimalis), and
tarpon {Mejjalops atlanticus). Along the northern Gulf Coast, juvenile brown and
white shrimp {Pmaeus spip .) and male blue crabs {Callinectes sapidus) may also move
into freshwater areas. These juveniles emigrate to deeper, more saline waters as
temperatures drop in the fall.

Birds

Of all wetland habitats, coastal freshwater marshes may support the largest and
most diverse populations of birds. W. E. Odum et al. (1984), working from a
number of studies, compiled a list of 280 species of birds that have been reported
from tidal freshwater marshes. They stated that although it is probably true that
this environment supports the greatest bird diversity of all marshes, the lack of
comparative quantitative data makes it difficult to test this hypothesis. Bird species
include: waterfowl (44 species); wading birds (15 species); rails and shorebirds (35
species); birds of prey (23 species); gulls, terns, kingfishers, and crows (20 species);
arboreal birds (90 species); and ground and shrub birds (53 species). A major reason
for the intense use of these marshes is the structural diversity of the vegetation
provided by broad-leaved plants, tall grasses, shrubs, and interspersed ponds.

Dabbling ducks (family Anatidae) and Canada geese actively select tidal freshwater
areas on their migratory flights from the North. They use the Atlantic Coast marshes in
the late fall and early spring, flying farther south during the cold winter months. Most
of these species winter in fresh coastal marshes of the northern Gulf of Mexico, but
some fly to South America. Their distribution in apparently similar marshes is variable;
some marshes support dense populations, others few birds. For example. Fuller et al.
( 1988 ) found extensive use of new Atchafalaya River delta marshes by many species of
ducks. Although the vegetation was dominated by arrowhead throughout the newly
created islands, duck populations were twice as dense in the western islands of the delta
compared to the east and central islands. On the central islands, ducks preferentially
selected stands of three -square sedge over arrowhead; on the western islands where
there was not three-square sedge, they frequented stands with mixed grass species
over arrowhead. The reason for this selectivity is unclear. In the Atchafalaya Delta, it
may be because the western islands remain fresh year-round, whereas the other islands
sometimes experience saltwater encroachment (Holm, 1998). The birds feed in fresh-
water marshes on the abundant seeds of annual grasses and sedges, the rhizomes of
perennial marsh plants, and also in adjacent agricultural fields. They are opportunistic
feeders, on the whole, ingesting from the available plant species. An analysis by Aber-
nethy (1986) suggests that many species that frequent the fresh marsh early in the
winter move seaward to salt marshes before beginning their northward migration in
the spring. The reason for this behavior pattern is not known, but Abernethy specu-
lated that the preferred foods of the freshwater marshes are depleted by early spring
and the birds move into salt marshes that have not been previously grazed.

294 Chapter 8 Tidal Marshes

The wood duck {Aix sponsa) is the only duck species that nests regularly in coastal
freshwater tidal marshes, although an occasional black duck {Anas rubripes) or mallard
{A. platyrhynchos) nest is found in Atlantic Coast marshes.

Wading birds are common residents of coastal freshwater marshes. They are
present year-round in Gulf Coast marshes but only during the summer along the
Adantic Coast. An exception is the great blue heron {Ardea herodias)^ which is seen
throughout the winter in the northern Atlantic states. Nesting colonies are common
throughout the southern marshes, and some species, such as green-backed herons
{Butorides striatus) and bitterns {Ixobrychus exilis and Botaurus Imtiginosus) ^ nest
along the Mid-Atiantic Coast. They feed on fish and benthic invertebrates, often
flying long distances each day from their nesting areas to fish.

Rails {Rallus spp .) and shorebirds, including the killdeer {Charadriusvocikms)^
sandpipers (Scolopacidae), and the American woodcock {Scolopax minor)^ are com-
mon in coastal freshwater marshes. They feed on benthic macroinvertebrates and
diverse seeds. Gulls (LarwTspp.), terns {Sterna spp .) ^ belted kingfishers {Ceryle alcyon),
and crows {Corvus spp.) are also common. Some are migratory; some are not. A num-
ber of birds of prey are seen hovering over freshwater marshes, including the northern
harrier {Circus cyaneus), the American kestrel {Falco sparverius), falcons {Falco spp.),
eagles, ospreys {Pandion haliaetus)^ owls (Tytonidae), vultures (Cathartidae), and the
loggerhead shrike {Lanius ludovicianus) . Swallow-tailed kites have been found to use
the lower tidal reach of forested wetlands for nesting (Sykes et ah, 1999). Arboreal
birds use the coastal freshwater marshes intensively during short periods of time on
their annual migrations. Flocks of tens of thousands of swallows (Hirundinidae) have
been reported over the upper Chesapeake freshwater marshes. Flycatchers (Tyran-
nidae) are also numerous. They often perch on trees bordering the marsh, darting out
into the marsh from time to time to capture insects. Although coastal marshes may
be used for only short periods of time by a migrating species, they may be impor-
tant temporary habitats. For example, the northern Gulf coastal marshes are the first
landfall for birds on their spring migration from South America. Often they reach this
coast in an exhausted state, and the availability of forested barrier islands for refuge
and marshes for feeding is critical to their survival.

Sparrows, finches (Fringillidae), juncos (Junco spp.), blackbirds (Icteridae), wrens
(Troglodytidae), and other ground and shrub birds are abundant residents of coastal
freshwater marshes. W. E. Odum et al. (1984) indicated that 10 species breed in
Mid -Atlantic Coast marshes, including the ring-necked pheasant {Phasianus colchicus)^
red-winged blackbird {A^elaius phoeniceus), American goldfinch {Carduelis tristis),
rufous-sided towhee {Pipilo erythrophthalmus)^ and a number of sparrows. The most
abundant are the red-winged blackbirds, dickcissels {Spiza americana ), and bobolinks
{DoUchonyx oryzivorus)^ which can move into and strip a wild rice marsh in a few days.

Amphibians and Reptiles

Although W. E. Odum et al. (1984) compiled a list of 102 species of amphibians
and reptiles that frequent coastal freshwater marshes along the Atlantic Coast, many

Tidal Freshwater Wetlands 295

are poorly understood ecologically, especially with respect to their dependence on this
type of habitat. None is specifically adapted for life in tidal tfeshwater marshes. Instead,
they are able to tolerate the special conditions of this environment. River turtles, the
most conspicuous members of this group, are abundant throughout the southeastern
United States. Three species of water snakes (Nerodia) are common. A^kistrodon pis-
civorus (the cottonmouth) is found south of the James River in Virginia. In the South,
especially along the Gulf Coast, the American alligator’s preferred habitat is the tidal
freshwater marsh. These large reptiles used to be listed as threatened or endangered,
but they have come back so strongly in most areas that currently they are harvested
legally (under strict control) in Louisiana and Florida. They nest along the banks of
coastal freshwater marshes, and the animal, identified by its high forehead and long
snout, is a common sight gliding along the surface of marsh streams.

Mammals

The mammals most closely associated with coastal freshwater marshes are all able to
get their total food requirements from the marsh, have fur coats that are more or
less impervious to water, and are able to nest (or hibernate, in northern areas) in the
marsh. These include the river otter (Lutm canadensis)^ muskrat ( Ondatra zibethicus),
nutria {Myocastor coypus), mink {Mustela vison), raccoon {Procyon lotor), marsh rabbit
{Silvilajjus palustris), and marsh rice rat (Oryzomys palustris). In addition, the opos-
sum {Didelphis virginiana) and white-tailed deer {Odocoileus virginianus) are locally
abundant. The nutria was introduced from South America some years ago and has
spread steadily in the Gulf Coast states and into Maryland, North Carolina, and Vir-
ginia. It is not likely to spread farther north because of its intolerance to cold, but
the South Atlantic marshes would seem to provide an ideal habitat. The nutria is
more vigorous than the muskrat and has displaced it from the freshwater marshes
in many parts of the northern Gulf As a result, muskrat density is highest in oligoha-
line marshes. The muskrat, for some reason, is not found in coastal Georgia and South
Carolina or in Florida, although it is abundant farther north along the Atlantic Coast.
Muskrat, nutria, and beaver ( Castor canadensis) can influence the development of a
marsh. The first two species destroy large amounts of vegetation with their feeding
habits (they prefer juicy rhizomes and uproot many plants when digging for them),
their nest building, and their underground passages. Beavers have been observed in
tidal freshwater marshes in Maryland and Virginia. Their influence on forested habi-
tats is well known, but their impact on tidal freshwater marshes needs to be studied
more closely.

Ecosystem Function

Primary Productivity

Many production estimates have been made for freshwater coastal marshes. Pro-
ductivity is generally high, usually falling in the range of 1,000 to 3,000 g yr“^
(Table 8.5). The large variability reported from different studies stems, in part, from

296 Chapter 8 Tidal Marshes

Table 8.5 Peak standing crop and annual net primary production (NPP) estimates for tidal
freshwater marsh associations in approximate order from highest to lowest productivity^

Vegetation Type'’

Peak Standing
Crop (g m-2)

Annual NPP
(g m-2 yr-i)

Extremely High Productivity

Spartina cynosuroides (big cordgrass)

2,311

—

Lythrum salicaria (spiked loosestrife)

1,616

2,100

Zizaniopsis miliacea (giant cutgrass)

1,039

2,048

Panicum hemitomon (maidencane)

1,160

2,000

Phragmites communis (common reed)

1,850

1,872

Moderate Productivity

Zizania aquatica (wild rice)

1,218

1,578

Amaranthus cannabinus (water hemp)

960

1,547

Typha sp. (cattail)

1,215

1,420

Bidens spp. (bur marigold)

1,017

1,340

Polygonum sp./Leers/a oryzoides (smartweed/rice cutgrass)

1,207

—

Ambrosia tirifida (giant ragweed)

1,205

Acorus calamus (sweet flag)

857

1,071

Sagittaria latifolia (duck potato)

432

1,071

Low Productivity

Peltandra virginica/Pontederia cordata (arrow arum/pickerelweed)

671

888

Hibiscus coccineus (rose mallow)

1,141

869

Nuphar adventa (spatterdock)

627

780

Rosa palustris (swamp rose)

699

—

Scirpus deltarum

—

523

Eleocharis baldwinii

130

—

^Values are means of 1 to 8 studies.

'’Designation indicates the dominant species in the association.

Sources: W. E. Odum et al. (1984); Sasser and Gosselink (1984); Visser (1989); White (1993); and Sasser
et al. (1995)

a lack of standardization of measurement techniques, but real differences can be
attributed to three factors:

1. Type of plant and its growth habit. Fresh coastal marshes, in contrast to saline
marshes, are floristically diverse, and productivity is determined, at least to
some degree, by genetic factors that regulate the species’ growth habits. Tall
perennial grasses, for example, appear to be more productive than
broad-leaved herbaceous species such as arrow arum and pickerelweed.

2. Tidal enerpiy. The stimulating effect of tides on production has been shown
for salt marshes and appears to be true for tidal freshwater marshes as well.

3. Other factors. Soil nutrients, grazing, parasites, and toxins are other factors
that can limit production in tidal freshwater marshes.

The elevation gradient across a fresh coastal marsh and the resulting differences
in vegetation and flooding patterns account for three broad zones of primary

Tidal Freshwater Wetlands 297

production. The low marsh bordering tidal creeks, dominated by broad-leaved
perennials, is characterized by apparently low production rates. Biomass peaks early
in the growing season. Turnover rates, however, are high, suggesting that annual
production may be much higher than can be determined from peak biomass. Much
of the production is stored in below- ground biomass (root : shoot » 1) in mature
marshes; this biomass is mostly rhizomes rather than fibrous roots. Decomposition is
rapid, the litter is swept from the marsh almost as fast as it forms, the soil is bare in
winter, and erosion rates are high. The parts of the high marsh dominated by perennial
grasses and other erect, tall species are characterized by the highest production rates
of freshwater species, and root : shoot ratios are approximately 1 . Because tidal energy
is not as strong and the plant material is not so easily decomposed, litter accumulates
on the soil surface, and little erosion occurs. The high-marsh mixed-annual associ-
ation typically reaches a large peak biomass late in the growing season. Most of the
production is above-ground (root : shoot <1), and litter accumulation is common.

Few estimates of primary productivity have been conducted in tidal freshwater
swamps; however, they are expected to benefit from the same tidal nutrient subsidy.
Relative elevation can have an effect of tree composition, geomorphology, flooding
duration, and tidal exchange — all of which may influence productivity. There is likely
a broad range of productivity and contribution for tidal freshwater wetlands. Tree
growth in forests near their downstream threshold can be stunted by higher salini-
ties and inundation frequency. Upstream, however, forests may be highly productive,
benefiting from the nutrient subsidy provided by the tides and diminished saltwa-
ter intrusion. Ozalp et al. (2007) found that above-ground net primary productivity
(ANPP) at a tidal forest along the lower Pee Dee River in South Carolina ranged
between 477 and 1,117 g nr^ yr^h Along the tidal reaches of the Pamunkey River,
Fowler (1987) found that forest ANPP at 1,230 g nr^ yr^\ with 40 percent of this
production from herbaceous plants. Along Lake Maurepas in southeastern Louisiana,
where flooding durations and depths have increased due to anthropogenic alterations,
Effler et al. (2007) reported that annual tree production of tidal swamps was low,
ranging between 220 and 700 g nr^ yr^h

Energy Flow

There are three major sources of organic carbon to tidal freshwater marshes. The
largest source is probably the vascular marsh vegetation, but organic material brought
from upstream (terrestrial carbon) may be significant, especially on large rivers and
where domestic sewage waters are present. Phytoplankton productivity is a largely
unknown quantity. Most of the organic energy flows through the detrital pool and is
distributed to benthic fauna and deposit-feeding omnivorous nekton. These groups
feed fish, mammals, and birds at higher trophic levels. The magnitude of the herbi-
vore food chain, in comparison to the detritus one, is poorly understood. Insects are
more abundant in fresh marshes than in salt marshes but most do not appear to be
herbivorous. Marsh mammals apparently can “eat out” significant areas of vegetation
(Evers et ah, 1998), but direct herbivory is probably small in comparison to the flow
of organic energy from destroyed vegetation into the detrital pool. Nevertheless, these

298 Chapter 8 Tidal Marshes

rodents may exert strong control on species composition and on primary production
(Evers et al., 1998). Herbivores also act in synergy with other stresses, for example,
saltwater intrusion, flooding, and fire (Taylor et al., 1994; Grace and Ford, 1996).

The phytoplankton-zooplankton-juvenile fish food chain in fresh marshes is of
interest because of its importance to humans. Zooplankton are an important dietary
component for a variety of larval, posdarval, and juvenile fish of commercial impor-
tance that are associated with tidal freshwater marshes (Van Engel and Joseph, 1968).

Birds are major seasonal or year-round consumers in all types of tidal freshwater
marshes. In addition, they move materials out of the system, processing it into guano,
which in some areas may be a significant source of nutrients. They also modify plant
composition and production by eat-outs (Smith and Odum, 1981).

Organic Import and Export

In mature tidal freshwater marshes, most organic production is decomposed to litter
and peat within the marsh system, and nutrients are extracted and recycled. Floating
tidal freshwater marshes may have even more closed cycles. Because they float, no sur-
face flows export or import organic material. This limits fluxes to subsurface dissolved
materials. The largest loss of organic energy in these mature marshes is probably to
deep peats in the case of anchored marshes or to an organic sludge layer under the
water column in floating marshes (Sasser et al., 1991). The magnitude of this loss was
measured as 145 to 1 50 g C yr*^ in a Gulf Goast freshwater marsh (Hatton, 1981).

Other losses of organic carbon from marshes occur through flushing from the
marsh surface, conversion to methane that escapes as a gas, and export as biomass
in the bodies of consumers that feed on the marsh. In highly reduced freshwater
sediments, where, in contrast to salt marshes, littie sulfur is available as an electron
acceptor, it is expected that methanogenesis from carbon dioxide and fermentation
should be a dominant pathway of respiratory energy flow. However, there is evidence
that macrophyte biomass may regulate methanogenesis. Neubauer et al. (2005) exam-
ined the anaerobic metabolism in soils from tidal (freshwater and salt) marshes along
the Patuxent River, Maryland. In the freshwater tidal marsh, they found that anaer-
obic metabolism was dominated by iron (111) reduction early in the growing season
(Fig. 8.13). Eater, when plant biomass declined, methanogenesis became the domi-
nant pathway for anaerobic metabolism. The authors attributed the enhanced ferric
iron (Fe'*'^) reduction to radial oxygen loss from plants during their peak growing
period that allow the replenishment of Fe'^^oxides back into the rhizosphere. Rates
of anaerobic metabolism were lower in the salt marsh and were dominated almost
equally by both Fe reduction and sulfate reduction when plants were most productive
and almost exclusively by sulfate reduction late in the growing season. The relationship
between plant biomass and iron reduction was less clear in the salt marsh (Fig. 8.13)
and may have been confounded by shifts in seasonal flooding.

Garbon dynamics in tidal freshwater forests can be variable and likely change along
the estuarine gradient. Tidal export of detritus matter is likely significant; however,
there is little information and few estimates available for these wetlands. Soil organic
matter tends to be high in these wetlands and is often linked to hydrology. A review of

Tidal freshwater marsh Salt marsh

■ Fe (III) reduction □ ’ reduction A methanogenesis

Figure 8.13 Seasonal changes in three anaerobic metabolism processes (iron reduction,
sulfate reduction, and methanogeneis) in Maryland tidal freshwater and salt marshes. In the
tidal freshwater marsh, iron reduction dominated early in the growing season (coinciding
with peak plant biomass) and methanogenesis later. In the salt marsh, rates were lower and
were dominated by iron and sulfate reduction early and by only sulfate reduction later in the
growing season. (After Neubauer et al., 2005)

•Q

Figure 8.14 Methane fluxes from static chambers and concurrent tidal water level changes
in a tidal freshwater swamp in White Oak River Estuary, North Carolina. The highest methane
emission occurred while water levels coincided with the soil surface. When water levels were
below the soil surface, methane oxidation increased in an aerobic surface layer, resulting in
decreased emission. When water levels exceeded the soil surface, methane emission was
probably reduced by a water diffusion barrier. (After Kelley et al., 1995)

299

300 Chapter 8 Tidal Marshes

reported soil conditions in the southeastern United States found that the percentage
of soil organic matter at the surface ranged from 9 to 77 percent with the highest
concentrations reported for blackwater rivers (Anderson and Lockaby, 2007). For
comparison, surface soils in a tidal freshwater shrub wetland in the Netherlands had
35 percent organic matter (Verhoeven et ah, 2001). At the lower river reaches where
waters can be more brackish, carbon mineralization and anaerobic metabolism may
alternate between sulfate reduction and methanogenesis, although iron reduction may
be important as well. During low tides, water levels in these swamps can drop below
the sediment surface, and the increased aerobic conditions can increase methane oxi-
dation and decrease emissions from these wetlands; during high tides, inflowing water
can both cool soil temperatures and provide a medium for methane oxidation, reduc-
ing methane generation then as well (Fig. 8.14). Consequently, methane flux rates in
tidal freshwater marshes and swamps tend to be lower than in comparable nontidal
freshwater marshes and swamps (Anderson and Lockaby, 2007).

Figure 8.15 Nitrogen budget for a 23-ha tidal freshwater marsh in coastai Massachusetts.
Pool sizes are in kmoles N, fluxes in kmoles N yr~^. (After Bowden et al., 1991)

References 301

Nutrient Budgets

In general, nutrient cycling and nutrient budgets in coastal freshwater wetlands appear
to be similar to salt marshes and mangrove swamps: They are fairly open systems that
have the capacity to act as long-term sinks, sources, or transformers of nutrients. Even
though these marshes generally are vigorously flooded by tides, they recycle a major
portion of the nitrogen requirements of the vegetation. Figure 8.15 illustrates nitro-
gen cycling in a tidal freshwater marsh near Boston, Massachusetts. In this budget,
most nutrient inputs are inorganic from the North River, and the marsh and river
nutrient cycles are mostly independent. Within the marsh itself, the major cycle is
from peat to ammonium-nitrogen to live plants. Some of the ammonium-nitrogen
is nitrified to nitrate-nitrogen and denitrified. Overall nitrate loss always exceeded
denitrification measurements by the acetylene block method, suggesting that other
sinks of nitrate, such as assimilatory nitrate reduction, may be important (Bowden
et ah, 1991). Peat mineralization is sufficient to satisfy the nitrogen demands of the
vegetation, and nearly all of the nitrogen flowing across the marsh from the adjacent
river is reexported. Mineralized litter and peat are conserved in the marsh by plant
uptake and by microbial and litter immobilization. Despite this closed mineral cycle,
the small uptake of nitrogen from the river may be important.

Recommended Readings

Day, J. W., B. C. Crump, W. M. Kemp, and A. Yahez-Arancibia, eds. 2013. Estuarine
Ecology. Hoboken, NJ: John Wiley & Sons.

Perillo, G., E. Wolanski, D. Gaboon, and M. Brinson. 2009. Coastal Wetlands: An
Inte^raded Ecosystem Approach. Amsterdam: Elsevier.

Tiner, R. W. 2013. Tidal Wetlands Primer: An Introduction to Their Ecology, Natural
History, Status, and Conservation. Amherst: University of Massachusetts Press.

References

Abernethy, R. K. 1986. Environmental Conditions and Waterfowl Use of a Backfilled
Pipeline Canal. M.S. thesis, Louisiana State University, Baton Rouge. 125 pp.
Adam, P. 1998. Australian saltmarshes: A review. In A. J. McGomb and J. A.
Davis, eds.. Wetlands for the Euture. Gleneagles Publishing, Adelaide, Australia,
pp. 287-295.

Anderson, G. J. and B. G. Lockaby. 2007. Soils and biogeochemistry of tidal fresh-
water forested wetlands. In W. H. Gonner, T. W. Doyle, and K. W. Krauss, eds.,
Ecolojjy of Tidal Ereshwater Eorested Wetlands of the Southeastern United States.
Springer, Inc., Dordrecht, Netherlands, pp. 65-88.

Antifinger, A. E., and E. L. Dunn. 1979. Seasonal patterns of GO2 and water vapor
exchange of three salt-marsh succulents. Oecolopfia (Berk) 43: 249-260.

Baldwin, A. H., K. L. McKee, and I. A. Mendelssohn. 1996. The influence of veg-
etation, salinity, and inundation on seed banks of oligohaline coastal marshes.
American Journal of Botany 83: 470-479.

302 Chapter 8 Tidal Marshes

Baldwin, A. H., M. S. Egnotovish, and E. Clarke. 2001. Hydrologic change and vege-
tation of tidal freshwater marshes: Field, greenhouse, and seed-bank experiments.
Wetlands 21: 519-531.

Beadle, L. C. 1974. The Inland Waters of Tropical Africa. Longman, London.

Bouchard, V. and J. C. Lefeuvre. 2000. Primary production and macro-detritus
dynamics in a European salt marsh: carbon and nitrogen budgets. Aquatic Botany
67: 23^2.

Bouchard, V., M. Tessier, F. Digaire, J. P. Vivier, L. Valery, J. C. Gloaguen, and
J. C. Lefeuvre. 2003. Sheep grazing as management tool in Western European
saltmarshes. C.R. Biologies 326: S148-S157.

Bowden, W. B., C. J. Vorosmarty, J. T. Morris, B. J. Peterson, J. E. Hobbie, P. A.
Steudler, and B. Moore. 1991. Transport and processing of nitrogen in a tidal
freshwater wetland. Water Resources Research 27: 389-408.

Bradley, P. M., and E. L. Dunn. 1989. Effects of sulfrde on the growth of three salt
marsh halophytes of the southeastern United States. American Journal of Botany
76: 1707-1713.

Caldwell, R., and Crow. 1992. A floristic and vegetation analysis of a tidal freshwater
marsh on the Merrimac River, West Newberry, MA. Rhodora 94: 63-97.

Campeau, S., H. R. Murkin, and R. D. Titman. 1994. Relative importance of algae
and emergent plant litter to freshwater marsh invertebrates. Canadian Journal of
Fisheries and Aquatic Sciences 51: 681-692.

Chapman, V. J. 1960. Salt Marshes and Salt Deserts of the World. Interscience, New
York. 392 pp.

Chapman, V. J. 1976. Coastal Vegetation., 2nd ed., Pergamon Press, Oxford, UK.
292 pp.

Chapman, V. J. 1977. Wet Coastal Ecosystems. Elsevier, Amsterdam, 428 pp.

Childers, D. L. 1994. Fifteen years of marsh flumes: a review of marsh-water column
interactions in southeastern USA estuaries. In W. J. Mitsch, ed.. Global Wetlands:
Old World and New. Elsevier, Amsterdam, pp. 277-293.

Childers, D. L., J. W. Day, Jr., and H. N. McKellar, Jr. 2000. Twenty more years of
marsh and estuarine flux studies: revisiting Nixon (1980). In M. P. Weinstein and
D. A. Kreeger, eds.. Concepts and Controversies in Tidal Marsh Ecology. Kluwer
Academic Publishers, Netherlands, pp. 389M:21.

Coleman, J. M., and H. H. Roberts. 1989. Deltaic coastal wetlands. Geolojjie en Mijn-
bouw 68: 1-24.

Crain, C. M., B. R. Silliman, S. L. Bertness, and M. D. Bertness. 2004. Physical and
biotic drivers of plant distribution across estuarine salinity gradients. Ecolojjy 85:
2539-2549.

Curco, A., C. Ibanez, and J. W. Day, Jr. 2002. Net primary production and decom-
position of salt marshes of the Ebro Delta (Catalonia, Spain). Estuaries 25:
309-324.

de la Cruz, A. A. 1974. Primary productivity of coastal marshes in Mississippi. Gulf
Research Reports A. 351-356.

References 303

Dreyer, G. D., and W. A. Niering, eds. 1995. Tidal Marshes of Lon^ Island Sound:
Ecology, History and Restoration. The Connecticut College Arboretum, Bulletin
No. 34, New London, CT.

Effler, R. S., G. P. Shaffer, S. S. Hoeppner, and R. A. Coyer. 2007. Ecology of the
Maurepas Swamp: Effects of salinity, nutrients, and insect defoliation. Soils and
biogeochemistry of tidal freshwater forested wetlands. In W. El. Conner, T. W.
Doyle, and K. W. Krauss, eds., Ecolojjy of Tidal Ereshwater Eorested Wetlands of the
Southeastern United States. Springer, Dordrecht, Netherlands, pp. 349-384.

Eisenlohr, W. S. 1972. Hydrologic Investigations of Prairie Potholes in North Dakota,
1958-1969, Professional Paper 585, U.S. Geological Survey, Washington, DC.

Evers, D. E., C. E. Sasser, J. G. Gosselink, D. A. Fuller, and J. M. Visser. 1998.
The impact of vertebrate herbivores on wetland vegetation in Atchafalaya Bay,
Louisiana. Estuaries 2\\ 1-13.

Field, D. W., A. Reyer, P. Genovese, and B. Shearer. 1991. Coastal wetlands of the
United States- An accounting of a valuable national resource. Strategic Assess-
ment Branch, Ocean Assessments Division, Office of Oceanography and Marine
Assessments, National Ocean Service, National Oceanic and Atmosphere Admin-
istration, Rockville, MD.

Fowler, K. 1987. Primary production and temporal variation in the macrophyte com-
munity of a tidal freshwater swamp. M.A. thesis, Virginia Institute of Marine
Science, College of William and Mary, Williamsburg, Virginia.

Fuller, D. A., G. W. Peterson, R. K. Abernethy, and M. A. LeBlanc. 1988. The distri-
bution and habitat use of waterfowl in Atchafalaya Bay, Louisiana. In C. E. Sasser
and D. A. Fuller, eds.. Vegetation and Waterfowl Use of Islands in Atchafalaya
Bay. Coastal Ecology Institute, Louisiana State University, Baton Rouge, LA,
pp' 73-103.

Funk, D. W., L. E. Noel, and A. H. Freedman. 2004. Environmental gradients, plant
distribution, and species richness in artic salt marsh near Prudhoe Bay, Alaska.
Wetlands Ecology and Mana£!ement 12: 215-233.

Gallagher, J. L. and F. C. Daiber. 1974. Primary production of edaphic algae com-
munities in a Delaware salt marsh. Limnolo£iy & Oceanojjraphy 19: 390-395.

Gosselink, J. G., Coleman, J. M., and R. E. Stewart, Jr. 1998. Coastal Louisiana. In
M. J. Mac, P. A. Opler, C. E. Puckett Haecker, and P. D. Doran, eds. Status and
Trends of the Nation’s Biological Resources, vol. 1 . U.S. Department of the Interior,
U.S. Geological Survey, Reston, VA, pp. 385-436.

Grace, J. B., and M. A. Ford. 1996. The potential impact of herbivores on the sus-
ceptibility of the marsh plant Sagittaria lancitblia to saltwater intrusion in coastal
wetlands. Estuaries 19: 13-20.

Greenberg, R., J. E. Maldonado, S. Droege, and M. V. McDonald. 2006. Tidal
marshes: A global perspective on the evolution and conservation of their
terrestrial vertebrates. BioScience 56: 675-685.

Hackney, C. T., G. B. Avery, L. A. Leonard, M. Posey, and T. Alphin. 2007. Bio-
logical, chemical, and physical characteristics of tidal freshwater swamp forests of

304 Chapter 8 Tidal Marshes

the lower Cape Fear River/Estuary, North Carolina. In W. H. Conner, T. W.
Doyle, and K. W. Krauss, eds.. Ecology of Tidal Freshwater Forested Wetlands of the
Southeastern United States. Springer, Inc., Dordrecht, Netherlands, pp. I83-22I.

Hall, S. L., and F. M. Fisher. 1985. Annual productivity and extracellular release of
dissolved organic compounds by the epibenthic algal community of a brackish
marsh. Journal ofPhycolojfy 21: 277-281.

Hall, J. V., W. E. Frayer, and B. O. When. 1994. Status of Alaska Wetlands. U.S. Fish
& Wildlife Service, Alaska Region, Anchorage, AK, 32 pp.

Hatton, R. S. 1981. Aspects of Marsh Accretion and Geochemistry: Barataria Basin,
Fa. Master’s thesis, Louisiana State University, Baton Rouge, LA.

Holm, G. O., Jr. 1998. Comparisons of the Atchafalaya and West Lake Delta salinity
incursions and the loss of Sagittaria latifolia Willd. M.S. thesis. Louisiana State
University, Baton Rouge, Louisiana.

Hopkinson, C. S. 1985. Shallow water benthic and pelagic metabolism: Evidence for
heterotrophy in the nearshore. Marine Biolojjy 87: 19-32.

Hopkinson, C. S., J. G. Gosselink, and F. T. Parronndo. 1980. Production of coastal
Louisiana marsh plants calculated from phenometric techniques. Ecolojjy 61:
I09I-I098.

Huffman, R. T., and R. E. Lonard. 1983. Successional patterns on floating vegetation
mats in a southwestern Arkansas bald cypress swamp. Castanea 48:73 78.

Ibanez, G., J. W. Day, Jr., and D. Pont. 1999. Primary production and decomposition
in wedands of the Rhone Delta, France: Interactive impacts of human modifica-
tions and relative sea level rise. Journal Coastal Research 15: 7I7-73I.

Ibanez, G., J. T. Morris, I. A. Mendelssohn, and J. W. Day, Jr. 2013. Goastal marshes.
In J. W. Day, Jr., B. G. Crump, W. M. Kemp, and A. Yanez-Arancibia, eds.. Estu-
arine Ecology, 2nd ed. Wiley-Blackwell, Hoboken, NJ, pp. I29-I63.

Junk, W. J. 1970. Investigations on the ecology and production biology of the “float-
ing meadows” (Paspalo echinochloetum) on the middle Amazon: I. The floating
vegetation and its ecology. Amazonia 2: 449-495.

Kelley C. A., C. S. Martens, and W. Ussier III. 1995. Methane dynamics across a
tidally influenced riverbank margin. Fimnolo^y & Oceanography 40: III2-II29.

Koch, M. S., I. A. Mendelssohn, and K. L. McKee. 1990. Mechanism for the hydro-
gen sulfide -induced growth limitation in wetland macrophytes. Fimnolo^y and
Oceanography 35: 399-408.

Kreeger, D. A., and R. I. E. Newell. 2000. Trophic complexity between producers and
invertebrate consumers in salt marshes. In M. P. Weinstein and D. A. Kreeger, eds..
Concepts and Controversies in Tidal Marsh Ecology. Kluwer Academic Publishers,
Netherlands, pp. 187-220.

Leek, M. A., and K. J. Graveline. 1979. The seed bank of a freshwater tidal marsh.
American Journal of Botany 66: I006-I0I5.

Leek, M. A., and R. L. Simpson. 1987. Seed bank of a freshwater tidal wetland:
turnover and relationship to vegetation change. American Journal of Botany 74:
360-370.

References 305

Lcck, M. A., and R. L. Simpson. 1995. Ten-year seed bank and vegetation dynamics
of a tidal freshwater marsh. American Journal of Botany 82: 1547-1557.

Lefeuvre, J. C., and R. F. Dame. 1994. Comparative studies of salt marsh processes in
the New and Old Worlds: An introduction. In W. J. Mitsch, ed., Global Wetlands:
Old World and New. Elsevier, Amsterdam, pp. 169-179.

Lefeuvre J. C., V. Bouchard, E. Feunteun, S. Grare, R Lafaille, and A. Radureau.
2000. European salt marshes diversity and functioning: The case study of
the Mont Saint-Michel Bay, France. Wetlands Ecology and Management 8:
147-161.

Lefeuvre, J. C., R Laffaille, E. Feunteun, V. Bouchard, and A. Radureau. 2003. Bio-
diversity in salt marshes: From patrimonial value to ecosystem functioning. The
case study of Mont-Saint-Michel Bay. C.R. Biologies 326: S125-S131.

MacIntyre, H. L., and J. J. Cullen. 1995. Fine-scale vertical resolution of chlorophyll
and photosynthetic parameters in shallow-water benthos. Marine Ecology Progress
Series 122: 227-237.

Mendelssohn, 1. A., and J. T. Morris. 2000. Eco -physiological controls on the pro-
ductivity of Spartina alterniflora Loisel. In M. R Weinstein and D. A. Kreeger, eds.
Concepts and Controversies in Tidal Marsh Ecology. Kluwer Academic Rublishers,
The Netherlands, p. 59-80.

Mitsch, W. J., J. G. Gosselink, C. J. Anderson, and L. Zhang. 2009. Wetland Ecosys-
tems. John Wiley 8c Sons, Inc., Hoboken, NJ, 296 pp.

Montague, C. L., and R. G. Wiegert. 1990. Salt marshes. In R. L. Myers and J. J.
Ewel, eds. Ecosystems ofElorida. University of Central Florida Rress, Orlando, FL,
pp. 481-516.

Morris, J. T, B. Kjerfve, and J. M., Dean. 1990. Dependence of estuarine productivity
on anomalies in mean sea level. Limnology & Oceanography 35: 926-930.

Morris, J. T, F. V. Sundareshwar, C. T. Nietch, B. Kjerfve, and D. R. Gaboon. 2002.
Response of coastal wetlands to rising sea level. Ecology 83: 2869-2877.

Newell, S. Y., and D. Rorter. 2000. Microbial secondary production from saltmarsh
grass shoots, and its known and potential fates. In M. R Weinstein and D. A.
Kreeger, eds.. Concepts and Controversies in Tidal Marsh Ecology. Kluwer Aca-
demic Rublishers, Dordrecht, Netherlands, pp. 159-185.

Neubauer, S. C., K. Givler, S. Valentine, and J. R. Megonigal. 2005. Seasonal patterns
and plant-mediated controls of subsurface wetland biogeochemistry. Ecology 86:
3334-3344.

Nixon, S. W. 1980. Between coastal marshes and coastal waters — a review of twenty
years of speculation and research on the role of salt marshes in estuarine produc-
tivity and water chemistry. In PR Hamilton and K. B. MacDonald, eds.. Estuarine
and Wetland Processes. Plenum, New York, pp. 437-525.

Odum, E. P. 1968. A research challenge: Evaluating the productivity of coastal and
estuarine water. Proceedings Second Sea Grant Conference., University of Rhode
Island, pp. 63-64.

306 Chapter 8 Tidal Marshes

Odum, E. P. 1980. The status of three ecosystem-level hypotheses regarding salt
marsh estuaries: tidal subsidy, outwelling, and detritus-based food chains.
In V. S. Kennedy, ed.. Estuarine Perspectives. Academic Press, New York,
pp. 485-495.

Odum, E. P. 2000. Tidal marshes as outwelling/ pulsing systems. In M. P. Weinstein
and D. A. Kreeger, eds.. International Symposium: Concepts and Controversies
in Tidal Marsh Ecology. Kluwer Academic Publishers, Dordrecht, Netherlands,
pp. 3-7.

Odum, W. E, T. J. Smith in, J. K. Hoover, and C. C. Mclvor. 1984. The Ecology of
Freshwater Marshes of the United States East Coast: A Community Profile. U.S. Fish
and Wildlife Service, FWS/OBS-87/17, Washington, DC, 177 pp.

Ozalp, M., W. H. Conner, and B. G. Lockaby. 2007. Above-ground productivity and
litter decomposition in a tidal freshwater forested wetland on Bull Island, SC,
USA. Forest Ecology and Management 2AE-. 31^3.

Pallis, M. 1915. The structural history of Plav: The floating fen of the delta of the
Danube. J. Finn. Soc. Bot. 43: 233-290.

Perry, J. E., and C. H. Hershner. 1999. Temporal changes in the vegetation pattern
in a tidal freshwater marsh. Wetlands 19: 90-99.

Petillon, J., F. Ysnel, A. Canard, and J. C. Lefeuvre. 2005. Impact of an invasive plant
(Ellymus athericus) on the conservation value of tidal salt marshes in western
France and implications for management: Responses of spider populations. Bio-
logical Conservation 126: 103-117.

Pfeiffer, W. J. , and R. G. Wiegert. 1981. Grazers on Spartina and their predators. In L.
R. Pomeroy and R. G. Wiegert, eds. The Ecology of a Salt Marsh. Springer-Verlag,
New York, pp. 87-112.

Pinckney, J., and R. G. Zingmark. 1993. Modeling the annual production ofintertidal
benthic microalgae in estuarine ecosystems. Journal of Phycology 29: 396M07.

Pomeroy, L. R. 1959. Algae productivity in salt marshes of Georgia. Limnology &
Oceanography 4: 386-397.

Pomeroy, L. R., W. M. Darley, E. L. Dunn, J. L. Gallagher, E. B. Haines, and D. M.
Witney. 1981. Primary production. In L. R. Pomeroy and R. G. Wiegert, eds..
Ecology of a Salt Marsh. Springer-Verlag, New York, pp. 39-67.

Rheinhardt, R. 1992. A multivariate analysis of vegetation patterns in tidal freshwater
swamps of lower Ghesapeake Bay, USA. Bulletin of the Torrey Botanical Club 119:
192-207.

Sasser, G. E. and J. G. Gosselink. 1984. Vegetation and primary production in a float-
ing wetland marsh in Louisiana. Aquatic Botany 20: 245-255.

Sasser, G. E., J. G. Gosselink, and G. P. Shaffer. 1991. Distribution of nitrogen
and phosphorus in a Louisiana freshwater floating marsh. Aquatic Botany 41:
317-331.

Sasser, G. E., J. M. Visser, D. E. Evers, and J. G. Gosselink. 1995. The role of envi-
ronmental variables on interannual variation in species composition and biomass

References 307

in a subtropical minerotrophic floating marsh. Canadian Journal of Botany 73:
413-424.

Sasser, C. E., J. G. Gossehnk, E. M. Swenson, G. M. Swarzenski, and N. G. Leibowitz.
1996. Vegetation, substrate and hydrology in floating marshes in the Mississippi
river delta plain wedands, USA. Ve^etatio 122: 129-142.

Silliman, B. R., H. van de Koppel, M. D. Bertness, L. E. Stanton, and 1. A.
Mendelssohn. 2005. Drought, snails, and large-scale die-off of southern U.S.
salt marshes. Science 310: 1803-1806.

Simpson, R. L., R. E. Good, M. A. Leek and D. F. Whigham. 1983. The ecology of
freshwater tidal wetlands. BioScience 33: 255-259.

Smith, T. J., and W. E. Odum. 1981. The effects of grazing by snow geese on coastal
salt marshes. Ecolojiy 62: 98-106.

Stout, J. R 1978. An Analysis of Annual Growth and Productivity ofjuncus roemeri-
anus Scheele and Spartina alterniflora Loisel in Coastal Alabama. Ph.D. disserta-
tion, University of Alabama, Tuscaloosa.

Sullivan, M. J., and G. A. Moncreiff. 1988. Primary production of edaphic algal com-
munities in a Mississippi salt marsh. Journal ofPhycolojjy 24: 49-58.

Sullivan, M. J., and G. A. Gurrin. 2000. Gommunity structure and functional dynamics
of benthic microalgae in salt marshes. In M. PP. Weinstein and D. A. Kreeger, eds..
Concepts and Controversies in Tidal Marsh Ecology. Kluwer Academic Publishers,
The Netherlands.

Swarzenski, G., E. M. Swenson, G. E. Sasser, and J. G. Gossehnk. 1991. Marsh mat
flotation in the Louisiana Delta Plain. Journal of Ecology 79: 999-1011.

Sykes, P W. Jr, G. B. Kepler, K. L. Litzenberger, H. R. Sansing, E. T. R. Lewis, and
J. S. Hatfleld. 1999. Density and habitat of breeding swallow-tailed kites in the
lower Suwannee ecosystem, Florida. Journal ofEield Ornithology 70: 321-336.

Taylor, K. L., J. B. Grace, G. R. Guntenspergen, and A. L. Foote. 1994. The interac-
tive effects of herbivory and Are on an oligohaline marsh. Little Lake, Louisiana,
USA. Wetlands 14: 82-87.

Teal, J. M. 1962. Energy flow in the salt marsh ecosystem of Georgia. Ecology 43:
614-624.

Teal, J. M. 1986. The Ecology of Regularly Flooded Salt Marshes of New England: A
Community Profile. U.S. Fish and Wildlife Service, Washington, DG, Biol. Repp.
85(7.4), 61pp.

Teal, J. M., and B. L. Howes. 2000. Salt marsh values: Retrospection from the end
of the century. In M. P Weinstein and D. A. Kreeger, eds.. Concepts and Con-
troversies in Tidal Marsh Ecology. Kluwer Academic Publishers, The Netherlands,
pp. 9-19.

Turner, R. E. 1977. Intertidal vegetation and commercial yields of penaeid shrimp.
American Fisheries Society Transactions 106: 411M:16.

Turner, R. E., W. Woo, and H. R. Jitts. 1979. Estuarine influences on a continental
shelf plankton community. Science 206: 218-220.

308 Chapter 8 Tidal Marshes

Tyler, P. A. 1976. Lagoon of Islands, Tasmani — Deathknell for a unique ecosystem.^
Biolq/jical Conservation 9: 1-11.

Van Engel, W. A., and E. B. Joseph. 1968. Characterization of Coastal and Estu-
arine Fish Nursery Grounds as Natural Communities . Final Report, Bureau of
Commercial Fisheries, Virginia Institute of Marine Science, Glocester Point, VA,
43 pp.

Valery, L., V. Bouchard, and J. C. Lefeuvre. 2004. Impact of the invasive native species
Elymus athericus on carbon pools in a salt marsh. Wetlands 24: 268-276.

Van Raalte, C. D. 1976. Production of epibenthic salt marsh algae: light and nutrient
limitation. Limnology and Oceanography 21: 862-872.

Verhoeven, J. T. A. 1986. Nutrient dynamics in minerotrophic peat mires. Aquatic
Botany 2S: 117-137.

Verhoeven, J. T. A., D. F. Whigham, R. van Logtestijn, and J. O’Neill. 2001. A com-
parative study of nitrogen and phosphorus cycling in tidal and non-tidal river
wetlands. Wetlands 2\: 219-222.

Visser, J. M. 1989. The Impact of Vertebrate Herbivores on the Primary Production
of Sapjittaria Marshes in the Wax Lake Delta, Atchafalaya Bay, Louisiana. Ph.D.
dissertation, Louisiana State University, Baton Rouge, 88 pp.

Visser, J. M., C. E. Sasser, R. G. Ghabreck, and R. G. Linscombe. 1998. Marsh veg-
etation types of the Mississippi River deltaic plain. Estuaries 21: 818-828.

Visser, J. M., G. E. Sasser, R. H. Ghabreck, and R. G. Linscombe. 1999. Long-term
vegetation change in Louisiana tidal marshes. Wetlands 19: 168-175.

Wadsworth, J. R, Jr. 1979. Duplin River Tidal System. Sapelo Island, Georgia. Map
reprinted in Jan. 1982 by the University of Georgia Marine Institute, Sapelo
Island, GA.

Watzin, M. G., and J. G. Gosselink. 1992. Coastal Wetlands of the United States.
Louisiana Sea Grant Gollege Program, Baton Rouge, LA, and U.S. Fish and
Wildlife Service, Lafayette, LA, 15 pp.

Werme, G. E. 1981. Resource Partitioninpi in the Salt Marsh Fish Community. Ph.D.
dissertation, Boston University, Boston, 126 pp.

Wharton, G. H., W. M. Kitchens, E. G. Pendleton, and T. W. Sipe. 1982. The ecology
of bottomland hardwood swamps of the Southeast: A community profile. U.S.
Fish and Wildlife Service, Biological Services Program FWS/OBS-8 1/37, 1 33 pp.

Wheeler, B. D. 1980. Plant communities of rich fen systems in England and Wales.
Journal of Ecology 68: 365-395.

Whigham, D. F., and R. L. Simpson. 1975. Ecological Studies of the Hamilton Marshes.
Progress report for the period June 1974-January 1975, Rider Gollege, Biology
Department, Lawrenceville, NJ.

Whigham, D. F., RL. Simpson, and K. Lee. 1980. The Effect ofSewajje Effluent on the
Structure and Function of a Freshwater Tidal Wetland Water Resources. Research
Institute Report, Rutgers University, New Brunswick, NJ, 160 pp.

White, D. A. 1993. Vascular plant community development on mudflats in the Mis-
sissippi River delta, Louisiana, USA. Aquatic Bot. 45: 171-194.

References 309

White, D. A., T. E. Weiss, J. M. Trapani, and L. B. Thien. 1978. Productivity and
decomposition of the dominant salt marsh plants in Louisiana. Ecology 59:
751-759.

Wiegert, R. G., and B. J. Freeman. 1990. Tidal Salt Marshes of the Southeast Atlantic
Coast: A Community Profile. U. S. Department of Interior, Fish and Wildlife Ser-
vice, Biological Report 85 (7.29), Washington, DC.

Wilson, J. O., R. Buchsbaum, 1. Valiela, and T. Swain. 1986. Decomposition in salt
marsh ecosystems: Phenolic dynamics during decay of litter of Spartina alterni-
flora. Marine Ecology Progress Series 29 \ 177-187.

Zedler, J. B. 1980. Algae mat productivity: Comparisons in a salt marsh. Estuaries 3:
122-131.

Zhu, R., Y. Liu, J. Ma, H. Xu, and L. Sun. 2008. Nitrous oxide flux to the atmosphere
from two coastal tundra wetlands in eastern Antarctica. Atmospheric Environment
42: 2437-2447.

Mangrove swamp

Chapter 9

Mangrove Swamps

Mangrove swamps replace salt marshes as the dominant coastal ecosystems in
subtropical and tropical repfions. An estimated 140,000 to 170,000 km? of
manpirove wetlands are found throupfhout the world. Manpfrove wetlands are
limited in the United States (where there are approximately 5,000km? of
manpiroves) mostly to Florida coastlines and to emerging mangrove swamps on
the Louisiana coastline. Mangrove wetlands have been classified according to
their hydrodynamics and topography as fringe mangroves, riverine mangroves,
basin mangroves, and dwarf or scrub mangroves. The dominant mangrove
plant species are known for several adaptations to the saline wetland
environment, including prop roots, pneumatophores, salt exclusion, salt
excretion, and the production of viviparous seedlings. Their productivity and
organic export are closely related to their hydrogeomorphic setting.

The coastal salt marsh of temperate middle and high latitudes gives way to its analog,
the mangrove swamp, in tropical and subtropical regions of the world. The mangrove
swamp is an association of halophytic trees, shrubs, and other plants growing in brack-
ish to saline tidal waters of tropical and subtropical coastlines. This coastal, forested
wetland (called a mangal by some researchers) is infamous for its impenetrable maze
of woody vegetation, its unconsolidated peat that seems to have no bottom, and its
many adaptations to the double stresses of flooding and salinity. The word mangrove
comes from the Portuguese word mangue?ov “tree” and the English word for “a
stand of trees” and refers to both the dominant trees and the entire plant community.

Many myths have surrounded the mangrove swamp. It was described at one time
or another in history as a haven for wild animals, a producer of fatal “mangrove root
gas,” and a wasteland of little or no value. Researchers, however, have established
the importance of mangrove swamps in exporting organic matter to adjacent coastal
food chains, in providing physical stability to certain shorelines to prevent erosion, in

311

312 Chapter 9 Mangrove Swamps

protecting inland areas from severe damage during hurricanes and tidal waves, and in
serving as sinks for nutrients and carbon. The extensive literature on the mangrove
swamp on a worldwide basis grows exponentially. This interest probably stems from
the worldwide scope of these ecosystems, the many unique features that they pos-
sess, and their role in climate change — both as being on the front line of sea-level
rise in the tropics and their high productivity that leads to significant carbon seques-
tration (popularly called “blue carbon”). Much of the early literature on mangroves
concerned floristic and structural topics. Beginning in the early 1970s, the focus was
on hydrogeomorphology and the functional aspects of mangrove swamps. Since that
time, a significant literature on ecophysiology, primary productivity, stressors, food
chains, and detritus dynamics of mangrove ecosystems has been produced, along with
new work on nutrient cycling, mangrove restoration, valuation of mangrove resources,
blue carbon sequestration, and responses of mangroves to sea level changes.

Geographical Extent

Mangrove swamps are found along tropical and subtropical coastlines throughout the
world, usually between 25° N and 25° S latitude (Fig. 9.1a). Their limit in the North-
ern Hemisphere generally ranges from 24° to 32° N latitude, depending on the local
climate and the southern limits of freezing weather. There are an estimated 138,000
to 170,000 km^ of mangrove swamps in the world (Giri et al., 2011; Twilley and Day,
2013; Krauss et al., 2014), with more than half of those swamps found in the latitudi-
nal belts between 0° and 10° (Fig. 9.1b). Mangroves are divided into two groups — the
Old World mangrove swamps and the New World and West African mangrove swamps.
Over 50 species of mangroves exist, and their distribution is thought to be related to
continental drift in the long term and possibly to transport by early humans in the short
term. The distribution of these species, however, is uneven. The swamps are particu-
larly dominant in the Indo-West Pacific region (part of the Old World group), where
they contain the greatest diversity of species. There are 36 species of mangroves in that
region, whereas there are only about 10 mangrove species in the Americas (Fig. 9.1c).
It has been argued, therefore, that the Indo-Malaysian region was the original cen-
ter of distribution for the mangrove species (Chapman, 1976). Certainly some of the
most intact mangrove forests in the world are found in Malaysia and in Micronesia, in
the small islands east of the Philippines in the western Pacific. Studies have illustrated
how important these mangrove swamps are to local economies in these regions (Ewel
et al., 1998; Cole et ah, 1999). Several mangrove species, not native to the Hawaiian
Archipelago despite its appropriate climate and coastal geomorphology, invaded the
islands in the early twentieth century and are now permanent fixtures on coastlines
there (Allen, 1998).

There is also a great deal of segregation between the mangrove vegetation found
in the Old World region and that found in the New World of the Americas and
West Africa. Two of the primary genera of mangrove trees, Rhizophora (red man-
grove) and Avicennia (black mangrove), contain separate species in the Old and New
Worlds, suggesting “that speciation is taking place independently in each region”
(Chapman, 1976).

Geographical Extent 313

e>, a

Figure 9.1 Distribution of mangrove wetlands (a) in the world and (b) by latitude, and (c)
mangrove species richness by lengitude. (After Chapman, 1977; Twilley et al., 1992; Ellison
et al. 1999)

Most of the mangrove swamps in the United States, estimated to cover over
5,000 km2, are found in Florida (see Table 8.1). The best development of mangroves
in Florida is along the southwest coast, where the Everglades and the Big Cypress
Swamp drain to the sea. Mangroves extend up to 30 km inland along water courses on
this coast. The area includes Florida’s Ten Thousand Islands, one of the largest man-
grove swamps in the world at 600 km^ . Because of development pressure, a significant
fraction of the original mangroves on these islands has been lost or altered. Patterson
(1986) reported that there was a loss of 24 percent of mangroves on one of the most
developed islands in this region, Marco Island, from 1952 to 1984. Mangroves are
now protected in Florida, and it is illegal to remove them from the shoreline.

Mangrove swamps are also common farther north along Florida’s coasts, north
of Cape Canaveral on the Atlantic coast and to Cedar Key on the Gulf of Mexico,

314 Chapter 9 Mangrove Swamps

where mixtures of mangrove and salt marsh vegetation appear. One species of man-
grove {Aviccnnia^erminans) is found in Louisiana and in the Laguna Madre of Texas,
and it has been spreading extensively for the last 40 years, a possible sign of climate
shift. Extensive mangrove swamps are also found throughout the Caribbean Islands,
including Puerto Rico. Lugo (1988) estimated that there were originally 120 km^ of
mangroves in Puerto Rico, although only half of those remained by 1975.

Geographic Limitation and Recent Expansion

The frequency and severity of frosts are the main factors that limit the extension of
mangroves beyond tropical and subtropical climes. For example, in the United States,
mangrove wetlands are found primarily along the Adantic and Gulf coasts of Florida
up to 27° to 29° N latitude, north of which they are replaced by salt marshes. The red
mangrove can survive temperatures as low as -2° to ^° C for 24 hours, whereas the
black mangrove can withstand several days at this temperature, allowing black man-
groves to extend farther north on Florida’s east coast than red mangroves (as far north
as 30° N). Similarly, Schaeffer-Novelli et al. (1990) described mangroves as extending
to 28° to 30° S latitude along the Brazilian coast. Three to four nights of a light frost
are sufficient to kill even the hardiest mangrove species. Lugo and Patterson-Zucca
(1977) showed that mangroves survived approximately five nonconsecutive days of
frost in January 1977 in Sea Horse Key Florida on the Gulf of Mexico shoreline (lati-
tude 29° N), but estimated that it would take 200 days for the forest to recover from
frost damage. They also hypothesized that soil salinity stress could modify frost stress
on mangroves, suggesting that the latitudinal limit of mangroves reflects a number of
stresses rather than one factor.

Gavanaugh et al. (2014) investigated recent changes of this latitudinal limit of
mangroves in Florida by reviewing 28 years of satellite imagery on the northern
extreme of mangrove extent on the Atlantic Ocean shoreline of Florida. They found a
poleward expansion of mangroves along this coastline from 1984 to 2011 (Fig. 9.2)
and a strong correlation between that expansion and a reduction in the frequency
of extreme cold events (days colder than -4°G). They concluded that this poleward
expansion of mangroves is not related to mean annual temperatures but rather
inversely related to the frequency of cold extreme temperatures, mosdy brought to
the Florida coasdine by polar fronts in the winter.

Hydrogeomorphology

There are several different types of mangrove wedands, each having a unique set of
topographic and hydrodynamic condidons. A classificadon scheme of five geomor-
phological setdngs where mangrove forests occur, as developed by Thom (1982),
includes systems dominated by waves, tides, and rivers or, most often, by combinadons
of these three energy sources. Like the coastal salt marsh, the mangrove swamp can
develop only where there is adequate protection from high-energy wave acdon. A
number of physiographic setdngs favor the protecdon of mangrove swamps, includ-
ing (1) protected shallow bays, (2) protected estuaries, (3) lagoons, (4) the leeward
sides of peninsulas and islands, (5 ) protected seaways, (6) behind spits, and (7) behind

Hydrogeomorphology 315

a) b)

latitude

Figure 9.2 (a) The Flerida, U.S., peninsula showing the long-term increase (soiid) or
long-term decrease (gray) in mangrove cover from 1980s to 2007-2011 for each 0.25° lati-
tude on the Atlantic Ocean coastline; (b) relationship between mangrove change (increase or
decrease) and iatitude. Vertical line on (b) indicates breakpoint of iatitude of 26.75° where
increase in mangrove cover begins. (After Cavanaugh et ai., 2014)

offshore shell or shingle islands. Unvegetated coastal and barrier dunes usually develop
where this protection does not exist, and mangroves are also often found behind these
dunes.

In addition to the required physical protection from wave action, the range and
duration of the flooding of tides exert a signihcant influence over the extent and
functioning of the mangrove swamp. Tides constitute an important subsidy for the
mangrove swamp, importing nutrients, aerating the soil water, and stabilizing soil
salinity. Salt water is important to the mangroves in eliminating competition from
freshwater species. Tides provide a subsidy for the movement and distribution of the
seeds of several mangrove species. They also circulate organic sediments in some fringe
mangroves for the benefit of hlter-feeding organisms, such as oysters, sponges, and
barnacles, and for deposit feeders, such as snails and crabs. Like salt marshes, mangrove
swamps are intertidal, although a large tidal range is not necessary. Most mangrove
wetlands are found in tidal ranges of 0.5 to 3 m or more. Mangrove tree species can also
tolerate a wide range of inundation frequencies. Rhizophora spp., the red mangrove,
is often found growing in continually flooded coastal waters below normal low tide.

At the other extreme, mangroves can be found several kilometers inland along
riverbanks where there is less tidal action. These mangroves depend on river discharge
and are nourished by river flooding in addition to infrequent tidal inundation and the
stability of groundwater and surface water levels near the coast.

Hydrodynamic Classification

The development of mangrove swamps is the result of topography, substrate, and
freshwater hydrology as well as tidal action. A classification of mangrove wetland
ecosystems according to their physical hydrologic conditions was developed in the

316 Chapter 9 Mangrove Swamps

Figure 9.3 Classification of mangrove wetlands according to four hydrogeomorphic classes
(and six types overall): (a) and (b) fringe mangroves; (c) riverine mangroves; (d) basin man-
groves; (e) and (f) scrub (dwarf) mangroves. (After Wharton et al., 1976; Lugo, 1980; Cintron
et al. 1985)

1970s by Ariel Lugo, Sam Snedaker, and others at the University of Florida. The four
major classes of mangrove wedands, based on their hydrogeomorphology, are shown
in Figure 9.3 and are discussed below.

1 . Fringe man£iroves. Fringe mangrove wetlands are found along protected shore-
lines, on narrow berms along the coastline or in wide expanses along gently sloping
beaches, and along some canals, rivers, and lagoons (Fig. 9.3a). If a berm is present,
the mangroves may be isolated from freshwater runoff and then have to depend com-
pletely on rainfall, the sea, and groundwater for their nutrient supply. A special case of
fringe mangroves are small islands and narrow extensions of larger and masses (spits)
that are “overwashed” on a daily basis during high tide. These are sometimes called

Hydrogeomorphology 317

overwash mangrove islands (Fig. 9.3b). The forests are dominated by the red man-
grove {Rhizophora) and a prop root system that obstructs the tidal flow and dissipates
wave energy during periods of heavy seas. Tidal velocities are high enough to wash
away most of the loose debris and leaf litter into the adjacent bay. The islands often
develop as concentric rings of tall mangroves around smaller mangroves and a per-
manent, usually hypersaline, pool of water. These wetlands are abundant in the Ten
Thousand Islands region of Florida and along the southern coast of Puerto Rico. They
are particularly sensitive to the effects of ocean pollution.

2. Riverine mangroves. Tall, productive riverine mangrove forests are found
along the edges of coastal rivers and creeks, often several miles inland from the coast
(Fig. 9.3c). These wedands may be dry for a considerable time, although the water
table is generally just below the surface. In Florida, freshwater input is greatest during
the wet summer season, causing the highest water levels and the lowest salinity in
the soils during that time. Riverine mangrove wedands export a significant amount
of organic matter because of their high producdvity. These wetlands are affected by
freshwater runoff from adjacent uplands and from water, sediments, and nutrients
delivered by the adjacent river. Hence they can be significandy affected by upstream
acdvity or stream alteradon. The combinadon of adequate fresh water and high
inputs of nutrients from both upland and estuarine sources causes these systems to be
generally very producdve, supporting large (16-26 m) mangrove trees. Salinity varies
but is usually lower than that of the other mangrove types described here. The flushing
of fresh water during wet seasons causes salts to be leached from the sediments.

3. Basin mangroves. Basin mangrove wedands occur in inland depressions, or
basins, often behind fringe mangrove wedands, and in drainage depressions where
water is stagnant or slowly flowing (Fig. 9.3d). These basins are often isolated from
all but the highest tides and yet remain flooded for long periods once dde water does
flood them. Because of the stagnant conditions and less frequent flushing by tides, soils
have high saUnides and low redox potendals. These wedands are often dominated by
black mangroves {Avicennia spp.) and white mangroves {Lapfuncularia spp.), and the
ground surface is often covered by pneumatophores from these trees.

These hydrogeomorphic classes of mangrove are broad categorizations, and
within these mangrove types, there are likely to be subtypes that can be defined by spe-
cific hydrologic conditions. Knight et al. (2008) characterized three basin forest sub-
types for A. marina forests using hydrology and forest structure data from the Coom-
babah Lake region in Southeast Queensland, Australia. These subtypes included a
“deep basin” (characterized by ~50 cm standing water, ~3 ddes yr”\ and mature tree
development), a “medium depth basin” (characterized by 1 5-30 cm standing water,
20-40 tides yr“\ and intermediate tree development), and a “shallow basin” (char-
acterized by 5-1 5 cm standing water, ~80 tides yr“^, and recent tree establishment).
Mangrove subtypes could be categorized in the other hydrogeomorphic classes as well.

4. Dwarf mangroves. There are several examples of isolated, low-productivity
scrub mangrove wetlands that are usually limited in productivity because of the lack
of nutrients or freshwater inflows. Dwarf mangrove wetlands are dominated by scat-
tered, small (often less than 2 m tall) mangrove trees growing in an environment that is

318 Chapter 9 Mangrove Swamps

probably nutrient poor (Fig. 9.3e). The nutrient-poor environment can be a sandy soil
or limestone marl. Hypersaline conditions and cold at the northern extremes of the
mangrove’s range can also produce “scrub,” or stressed mangrove trees, in riverine,
fringe, or basin wedands. True dwarf mangrove wetlands, however, are found in the
coastal fringe of the Everglades and the Florida Keys and along the northeastern coast
of Puerto Rico. Some of these wetlands in the Everglades are inundated by seawater
only during spring tides or storm surges and are often flooded by freshwater runoff
in the rainy season. These types of wetlands are actually an intermix of small red man-
grove trees with marsh vegetation such as sawgrass ( Cladium jamaicmse) and rush
(Juncus roemarianus) . Hammock mangrove wetlands also occur as isolated, slighdy
raised tree islands in the coastal fringe of the Florida Everglades and have character-
istics of both basin and scrub mangroves. They are slightly raised as a result of the
buildup of peat in what was once a slight depression in the landscape (Fig. 9.3f). The
peat has accumulated from many years of mangrove productivity, actually raising the
surface from 5 to 10 cm above the surrounding landscape.

Soils and Salinity

Soil salinity in mangrove ecosystems varies from season to season and with mangrove
type (Table 9.1). In riverine mangrove systems, the soil salinity is less than that of
normal seawater because of the influx of fresh water. In basin mangroves, however,
salinity can be well above that of seawater because of evaporative losses (>50 ppt).
Noted Florida naturalist John Henry Davis (1940) summarized four major points
about salinity in mangrove wedands from his studies in Florida years ago that still
hold today:

1. There is a wide annual variation in salinity in mangrove wetlands.

2. Saltwater is not necessary for the survival of any mangrove species but only
gives mangroves a competitive advantage over salt-intolerant species.

Table 9.1 Soil salinity ranges for major
mangrove types

Hydrodynamic Type

Soii saiinity, ppt

Fringe mangroves

Avicennia

Rhizophora

Riverine mangroves

10-20**

Basin mangroves

Avicennia

>50

Laguncuiaria

iow saiinity

mixed forest

30-40

^Higher in dry season when less freshwater
streamflow is available.

Source: Cintron et al., 1985

Vegetation 319

3. Salinity is usually higher and fluctuates less in interstitial soil water than in the
surface water of mangroves.

4. Saline conditions in the soil extend farther inland than normal high tide
because of the slight relief, which prevents rapid leaching.

Seasonal oscillations in salinity in mangrove wetlands are a function of the height
and duration of tides, the seasonality and intensity of rainfall, and the seasonality and
amount of tfesh water that enters the mangrove wetlands through rivers, creeks, and
runoff In Florida, summer wet-season convective storms and associated freshwater
flow in streams and rivers as well as an occasional hurricane in the late summer or early
fall lead to the dilution of saltwater and the lowest salinity concentrations. Salinity is
generally highest during the dry season, which occurs in the winter and early spring.

Soil Acidity

Mangrove soils are often acidic, although, in the presence of carbonate, as is often
the case in south Florida, the soil pore water can be close to neutrality. The soils
are often highly reduced, with redox potentials ranging from -100 to M:00mV. The
highly reduced conditions and the subsequent accumulation of reduced sulfldes in
mangrove soils cause extremely acidic soils in many mangrove areas. Dent (1986,
1992) reported a measured accumulation of 10 kg S m“^ of sediment per 100 years
in mangroves. When these soils are drained and aerated for conversion to agricultural
land, the reduced sulfldes, generally stored as pyrites, oxidize to sulfuric acid, produc-
ing what are known as cat clays. These highly acidic soils make traditional agriculture
difficult and are one of the reasons that, when mangrove swamps are converted to
fishponds, the ponds have a short lifetime before they are abandoned. Dent (1992)
argued that the “dereclamation” of some previously “reclaimed” marginal coastal soils
back to mangroves and salt marshes may be the best strategy for these acidic soils.

Vegetation

As is evident in coastal salt marshes, the stresses of waterlogged soils and salinity lead
to a relatively simple flora in most mangrove wetlands, particularly when compared
to their upland neighboring ecosystem, the tropical rain forest. There are more than
50 species of mangroves throughout the world (Stewart and Popp, 1987; Twilley
and Day, 2013), representing 12 genera in 8 families. Fewer than 10 species of man-
groves are found in the New World, and only 3 species are dominant in the south
Florida mangrove swamps — the red mangrove {Rhizophom mcm^le L. ), the black man-
grove {Avicennia ^erminansR. ^ also named A. nitida- Jacq.), and the white mangrove
{Lapfuncularia racemosa L. Gaertn.). Buttonwood {Conocarpus erecta L.), although
strictly not a mangrove, is occasionally found growing in association with mangroves
or in the transition zone between the mangrove wetlands and the drier uplands. Each
of the hydrologic types of mangrove wetlands described previously is dominated by
different associations of mangrove plants. Fringe mangrove wetlands are dominated

320 Chapter 9 Mangrove Swamps

Table 9.2 Structural characteristics of canopy vegetation for major mangrove types^

Hydrodynamic

Type

Number
of Tree
Species

Number of Trees (#/(ha)

Basal Area (m^/ha)

Stand
Height (m)

Aboveground

Biomass

(kg/m^)

>2.5 cm dbh

>10cm dbh

>2. 5cm dbh

>10 cm dbh

Fringe mangroves

1.7 ±0.1 (33)

4005 ± 642 (33)

852 ± 115 (31)

22.2 ± 1.5 (33)

14.6 ± 1.9 (31)

13.3 ± 2.6 (32)

0.8-15.9 (8)

Riverine

1.9 ± 0.1 (36)

1979 ± 209 (28)

661 ±71 (32)

30.4 ±3.5 (5)

32.6 ± 4.7 (32)

21.2 ± 4.8 (26)

1.6-28.7 (8)

mangroves
Basin mangroves

2.3 ±0.1 (31)

3599 ± 400 (31)

573 ± 102 (21)

18.5 ± 1.6 (31)

10.6 ± 2.2 (21)

9.0 ± 0.7 (31)

—

®Data are based on mangrove sites in Florida, Mexico, Puerto Rico, Brazil, Costa Rica, Panama, and Ecuador. Values are the
average ± standard error (number of observations) except for above-ground biomass, which is the range (number of observations).
Source'. Cintron et al. (1985)

by red mangroves (Rhizophom) that contain abundant and dense prop roots, partic-
ularly along the edges that face the open sea. Riverine mangrove wetlands are also
numerically dominated by red mangroves, although they are straight trunked and
have relatively few, short prop roots. Black [Apicmnia spp.) and white [La^uncu-
laria spp.) mangroves also frequently grow in these wetlands. Basin wetlands support
all three species of mangroves, although black mangroves are the most common in
basin swamps and hammock wetiands are mostly composed of red mangroves. Scrub
mangrove wetlands are typically dominated by widely spaced, short (less than 2 m tall)
red or black mangroves.

A comparison of the structural characteristics of the major hydrodynamic types
of mangrove wedands is provided in Table 9.2. These data were compiled from more
than 100 mangrove research sites throughout the New World. Fringe mangroves
generally have a greater density of large trees (>10cm diameter at breast height
(dbh)) compared to riverine and basin mangroves. Riverine wedands, however,
have the largest trees and, hence, a much greater basal area and tree height than
do fringe or basin mangroves. The biomass of riverine mangroves is generally the
highest, although data are difficult to compare because of the different methods and
sample sizes used in various observations. Cintron et al. (1985) reported a range
of aboveground biomass for the Florida mangroves of 9 to 17kg/m^ for riverine
mangroves and 0.8 to 15 kg/m^ for fringe mangroves. Single measures of 0.8 kg/m^
for a dwarf mangrove wedand and 9.8kg/m^ for a hammock mangrove (both in
Florida) were also reported.

Zonation

In trying to understand the vegetarion of mangrove wetlands, most early researchers
were concerned with describing plant zonation and successional patterns. Some
attempts were made to equate the plant zonation found in mangrove wedands with
successional seres, but Lugo (1980) warned that “zonarion does not necessarily
recapitulate succession because a zone may be a climax maintained by a steady
or recurrent environmental condirion.” J. H. Davis (1940) is generally credited
with the best early descriprion of plant zonarion in Florida mangrove swamps,
especially in fringe and basin mangrove wedands (Fig. 9.4). He hypothesized that the

Vegetation 321

Uplands Conocarpus Avicennia Rhizophora

transition zone zone zone

buttonwr

viviparous

sellings

black manqi

roves ■'
levee

pneumatophbres of marl

^High Tide
prop roots
^ Low Tide

‘ underlying rock:

Figure 9.4 Classic zonation pattern of Florida mangrove swamp with illustrations of man-
grove adaptations, such as prop roots, viviparous seedlings, and pneumatophores.

entire ecosystem was accumulating sediments and was therefore migrating seaward.
Typically, Rhizophora mangle is found in the lowest zone, with seedlings and small
trees sprouting even below the mean low tide in marl soils. Above the low-tide
level but well within the intertidal zone, full-grown Rhizophora with well-developed
prop roots predominate. There tree height is approximately 10 m. Behind these red
mangrove zones and the natural levee that often forms in fringe mangrove wetlands,
basin mangrove wetlands, dominated by black mangroves {Avicennia) with numerous
pneumatophores, are found. Flooding occurs only during high tides. Buttonwood
( Conocarpus erecta) often forms a transition between the mangrove zones and upland
ecosystems. Flooding occurs there only during spring tides or during storm surges,
and soils are often brackish to saline.

Thibodeau and Nickerson (1986) suggested that red mangroves have a much
lower ability to tolerate high sulfides typical of extremely reduced conditions than do
black mangroves, and so red mangroves occur in regions that are frequently flushed
by tides, whereas black mangroves are found in isolated basin settings where strongly
reduced substrate containing high sulfides are found and pneumatophores can be of
the greatest use (see the following section).

Mangroves as Land Builders?

The zonation of plants in mangrove wetlands led some researchers (e.g., J. H. Davis,
1940) to speculate that each zone is a step in an autogenic successional process that
leads to freshwater wedands and, eventually, to tropical upland forests or pine forests.
Other researchers, led by Egler (1952), considered each zone to be controlled by

322 Chapter 9 Mangrove Swamps

its physical environment to the point that it is in a steady state or at least a state of
arrested succession (allogenic succession). For example, with a rising sea level, the
mangrove zones migrate inland; during periods of decreasing sea level, the mangrove
zones move seaward. Egler thought that the impact of fire and hurricanes made con-
ventional succession impossible in the mangroves of Florida. Another theory, advanced
by Chapman (1976), is that mangrove succession may be a combination of both auto-
genic and allogenic strategies, or a “succession of successions.” If that is the case,
successional stages could be repeated a number of times before the next successional
level is attained.

Lugo (1980) reviewed mangrove succession in light of E. P. Odum’s criteria
(1969; see Chapter 7: Wetland Vegetation and Succession) and found that, except for
mangroves on accreting coastlines, traditional successional criteria do not apply. He
concluded that mangroves are true steady-state systems in the sense that they are the
optimal and self-maintaining ecosystems in low-energy tropical saline environments.
In such a situation, high rates of mortality, dispersal, germination, and growth are the
necessary tools of survival. Unfortunately, these attributes could lead many to identify
mangroves as successional systems.

It is no longer accepted dogma that mangrove wetlands are “land builders” that
are gradually encroaching on the sea, as was suggested by J. H. Davis (1940). Yet Lee
et al. (2014) state that “few of Davis’s critics offered contradictory data as extensively
detailed or impressive as those presented in his [Davis’s] classic works.” In many cases,
mangrove vegetation plays a passive role in the accumulation of sediments, and the
vegetation usually follows, not leads, the land building that is caused by current and
tidal energies. It is only after the substrate has been established that the vegetation
contributes to land building by slowing erosion and by increasing the rate of sediment
accretion.

The mangrove’s successional dynamics appear to involve a combination of ( 1 ) peat
accumulation balanced by tidal export, fire, and hurricanes over years and decades;
and (2) advancement or retreat of zones according to the fall or rise of sea level over
centuries. Some researchers (Alongi, 2008; Lee et al., 2014) refer to mangroves as
“land stabilizers” rather than “land builders.” When peat accumulation is added to
the stabilization of marine and riverine sediments, some mangroves can have both
horizontal movement as well as vertical movement upward (Fig. 9.5). As Figure 9.5
illustrates, with climate changes in sea level, temperature, atmospheric carbon dioxide
and rainfall all at the same time, the stability and/or expansion of mangroves is a
complicated matter.

Mangrove Adaptations

Mangrove vegetation, particularly the dominant trees, has several adaptations that
allow it to survive in an environment of high salinity, occasional harsh weather, and
anoxic soil conditions. (Chapter 7 gives an overview of wetland plant adaptations.)
These physiological and morphological adaptations have been ofinterest to researchers
and are among some of the most distinguishing features that the laypeople notice
when first viewing these wetlands. Some of the morphological adaptations are shown

Vegetation 323

temperature

atmosphertc (CO^

rainfall

marsh/
salt flat

i^itand

mangroves

sea-level

change

plant production

sedimentation

sod

elevation

physical

compaction

organic matter
decomposition

deep subsidence and tectonics

Figure 9.5 Model showing how climate change in sea level, temperature, atmospheric car-
bon dioxide (CO2), and rainfall, along with sedimentation, soil erosion, and tectonic uplift,
all influence the horizontal and vertical movement of coastal mangroves and possible land
building. (After Lee et al., 2014)

in Figure 9.6. Overall, physiological and morphological adaptations of mangroves
include (1) salinity control; (2) prop and drop roots, pneumatophores, and lenticles;
and (3) viviparous seedlings.

Salinity Control

Mangroves are facultative halophytes; that is, they do not require saltwater for growth
but are able to tolerate high salinity and thus outcompete vascular plants that do not
have this salt tolerance. The ability of mangroves to live in saline soils depends on
their ability to control the concentration of salt in their tissues. In this respect, man-
groves are similar to other halophytes. Mangroves have the ability both to prevent
salt from entering the plant at the roots {salt exclusion) and to excrete salt from the
leaves {salt secretion). Salt exclusion at the roots is thought to be a result of reverse
osmosis, which causes the roots to absorb only freshwater from saltwater. The root
cell membranes of mangroves species of Rhizophora, Avicennia, and Lajjuncularia,
among others, may act as ultrafilters that exclude salt ions. Water is drawn into the
root through the filtering membrane by the negative pressure in the xylem developed
through transpiration at the leaves; this action counteracts the osmotic pressure caused
by the solutions in the external root medium. There are also a number of mangrove
species (e.g., Avicennia and La^nuncularia) that have salt-secreting glands on the

Figure 9.6 Adaptations of mangroves, inciuding (a) prop roots of red mangroves
zophora) in south Florida, (b) drop roots of mangroves in western Costa Rica, (c) pneu-
matophores of black mangroves (Avicennia) in southwest Florida, (d) viviparous germinated
seedlings hanging in red mangrove canopy, and (e) a red mangrove seedling fioating verti-
cally in the water column, perhaps kilometers from where it fell into the water. (Photos by
W. J. Mitsch)

324

Vegetation 325

leaves to rid the plant of excess salt. The solutions that are secreted often contain high
concentrations of sodium chloride (NaCl) and salt crystals may form on the leaves.
Another possible way in which mangroves discharge salt is through leaf fall, although
the importance of this method is questioned.. This leaf fall may be significant because
mangroves produce essentially two crops of leaves per year.

Prop Roots and Pneumatophores

Some of the most notable features of mangrove wetlands are the prop roots and drop
rootsofthc red mangrove (Fig. 9.6a,b) and the numerous, sm-A\ pneumatophores o? the
black mangrove [Avicennia) (reaching 10-20 cm above the sediments and sometimes
considerably more; Fig. 9.6c). At a distance, drop roots resemble prop roots; they
extend from branches and other upper parts of the stem directly down to the ground,
hovering above or rooting only a few centimeters into the sediments. They adsorb
freshwater from rainfall into the plant, an especially important function in seasonally
dry climates. Oxygen enters the plant through small pores, called lenticles^ which are
found on both pneumatophores and prop and drop roots. When lenticels are exposed
to the atmosphere during low tide, oxygen is absorbed from the air, and some of
it is transported to and diffuses out of the roots through a system of aerenchyma
tissue. This maintains an aerobic microlayer around the root system. When prop roots
or pneumatophores of mangroves are continuously flooded by stabilizing the water
levels, those mangroves that have submerged pneumatophores or prop roots soon die.

In an interesting experiment to determine the importance of oxygen transport
from the aerial organs to the sediments, Thibodeau and Nickerson (1986) “capped”
with plastic tubing the pneumatophores of Avicennia^erminans in a fringe mangrove
forest in the Bahamas. They observed that the aerobic zone surrounding the roots
was reduced in the area by capping, indicating that the pneumatophores help the
plant produce an oxidized rhizosphere. They also found that the greater the number
of pneumatophores present in a given area, the more oxidized the soil. They described
the relationship as

£h = -307 + l.lpd (9.1)

where

= redox potential (mV)

pd = pneumatophore density (number per 0.25 m^)

Viviparous Seedlings

Red mangroves (and related genera in other parts of the world) have seeds that germi-
nate while they are still in the parent tree; a long, cigar-shaped hypocotyl {viviparous
seedling) develops while hanging from the tree (Fig. 9.6d). This is apparently an
adaptation for seedling success where shallow anaerobic water and sediments would
otherwise inhibit germination. The seedling (or propagule) eventually falls and often
will root if it lands on sediments or will float and drift in currents and tides if it falls into
the sea. After a time, if the floating seedling becomes stranded and the water is shallow
enough, it will attach to the sediments and root. Often the seedling becomes heav-
ier with time, rights itself to a vertical position in the water column (Fig. 9.6e), and
develops roots if the water is shallow. It is not well understood whether contact with

326 Chapter 9 Mangrove Swamps

the sediments stimulates root growth or if the soil contains some chemical compound
that promotes root development. The value of the floating seedlings for mangrove
dispersal and for invasion of newly exposed substrate is obvious. Rabinowitz (1978)
reported that the obligate dispersal time (the time required during propagule disper-
sal for germination to be completed) was 40 days for the red mangrove and 14 days
for the black mangrove propagules. She also estimated that red and black mangrove
propagules could survive floating in the water for 110 and 35 days, respectively.

Consumers

W. E. Odum et al. (1982) reported the following data from the literature describ-
ing faunal use of mangroves in Florida in terms of the number of species: 220 fish;
181 birds, including 18 wading birds, 29 water birds, 20 birds of prey, and 71 arboreal
birds; 24 reptiles and amphibians; and 18 mammals. In general, a wide diversity of ani-
mals is found in mangrove wedands; their distribution sometimes parallels the plant
zonation described previously. Many of the animals that are found in mangrove wet-
lands are filter feeders or detritivores, and the wetlands are just as important as a shelter
for most of the resident animals as they are a source of food. Some of the important
filter feeders found in Florida mangroves include barnacles {Balcmus eburneus) ^ coon
oysters (Ostreafrons), and the eastern oyster {Crassostrea vir^inica). These organisms
often attach themselves to the stems and prop roots of the mangroves within the
intertidal zone, filtering organic matter from the water during high tide.

Crabs are among the most important animal species in mangrove wetlands around
the world, and they appear to play an important role in maintaining biodiversity in
mangrove ecosystems. Taking into account the role that they have in seedling survival,
carbon cycling, sediment microtopography, and soil chemistry. Smith et al. (1991)
suggested that crabs are the keystone species of the mangrove ecosystem. They burrow
in the sediments, prey on mangrove seedlings, facilitate litter decomposition, and are
the key transfer organism for converting detrital energy to wading birds and fish in
the mangrove forest itself and to offshore estuarine systems. Mangroves around the
world are dominated by 6 of the 30 families of Brachyura that collectively make up
about 127 species. Uca and Sesarma are the most abundant crab genera in mangrove
wetlands. Uca spp. (fiddler crabs) are particularly abundant in mangrove wetlands in
Florida, living on the prop roots and high ground during high water and burrowing
in the sediments during low tide. Sesarma (sesarmids) is the most abundant genus
of crabs in the world, with dozens of species in the Indo-Malaysian and East African
mangroves and many fewer in tropical America.

One of the most significant ways in which crabs may influence the distribution
of mangroves is by selective predation of mangrove propagules. Yet this effect is not
common to all mangrove wetlands. In comparing the effect of this predation on man-
groves around the world. Smith et al. (1989) found that crabs consumed up to 75
percent of the mangrove propagules in Australian mangrove swamps but very litde of
the litterfall or propagules in Panamanian and Florida mangrove swamps.

The role of crabs in leaf litter removal (burial and consumption) has been illus-
trated in a number of studies. Robertson and Daniel (1989) estimated that in some

Ecosystem Function 327

mangrove forests, leaf processing by sesarmid crabs alone was over 75 times the rate
as microbial degradation and that crabs removed more that 70 percent of the litter
of the high-intertidal Ceriops and Bru^uiem mangroves in tropical Australia. Smith
et al. (1991) developed an experiment where crabs were removed from experimental
plots in Rhizophora forests. They found that the removal of crabs caused significantly
higher concentrations of sulfide and ammonium in the mangrove soils, due primarily
to the absence of burrowing, which oxygenates the soil.

Many other invertebrates, including snails, sponges, flatworms, annelid worms,
anemone, mussels, sea urchins, and tunicates, are found growing on roots and stems
in and above the intertidal zone. Wading birds frequendy found in Florida man-
groves include the wood stork {Mycteria americana), white ibis {Eudocimus albus),
roseate spoonbill {Ajaia ajaja), cormorant {Phalacrocorx spp.), brown pelican {Pel-
icanus occidentalis), egrets, and herons. Vertebrates that inhabit mangrove swamps
include alligators, crocodiles, turdes, bears, wildcats, pumps, and rats.

Ecosystem Function

Certain fimcdons of mangroves, such as primary productivity, organic export, and out-
welhng, have been studied extensively. A picture of the dynamics of mangrove wedands
has emerged from several key studies. These studies have demonstrated the importance
of the physical condidons of tides, salinity, and nutrients to these wedands and have
shown where natural and human-induced stresses have caused the most effect.

Primary Productivity

Based on a global assessment of mangrove producdvity. Bouillon et al. (2008) con-
servatively esdmated worldwide mangrove producdvity at 218 ± 72 teragram (Tg =
10^^ g)“C yr“^ and added that over half of the carbon is unaccounted for, based on
various estimates of mangrove carbon sinks (organic export, burial, and minerahza-
don). A wide range of productivity has been measured in mangrove wedands due to
the wide variety of hydrodynamic and chemical conditions encountered. Table 9.3
presents a balance of carbon flow in several fringe and basin mangrove swamps in
Florida and Puerto Rico. Net primary producdvity ranges from 570 to 2,700 g-C
yr“^ (equivalent to 1,200 to 6,000 g-dry wt )■ Gross and net primary produc-

tivity are highest in riverine mangrove wedands, lower in fringe mangrove wedands,
and lowest in basin mangrove wetlands. The highest producdvity in riverine mangrove
wetlands is due to the greater influence of nutrient loading and freshwater turnover
at the riverine site .

The important factors that control mangrove function in general and primary
productivity in particular are: (1) tides and storm surges; (2) freshwater discharge;
(3) parent substrate; and (4) water-soil chemistry, including salinity, nutrients, and
turbidity. These factors are not mutually exclusive, as rides influence water chem-
istry and hence producdvity by transpordng oxygen to the root system, by removing
the buildup of toxic materials and salt from the soil water, by controlling the rate

328 Chapter 9 Mangrove Swamps

Table 9.3 Mass balance of carbon flow (g-C m~^ yi^^) in mangrove forests in Florida and
Puerto Rico

Rookery Bay, Florida® Puerto Fahkahatchee Bay, Florida®

Rico

Fringe Basin Fringe*’ Basin Fringe Fringe

Gross primary productivity

Canopy

Algae

Totai

Respiration (plants)

Leaves, stems
Roots, above-ground
Roots, below-ground

Total plant respiration
Net primary production
Growth
Litterfaii

Respiration (heterotrophs)
Respiration (totai)

Export

Net ecosystem production

Burial

Growth

2,055

3,292

3,004

402

276

2,457

3,318

3,280

671

2,022

1,967

197

741

693

2,219

2,708

1,764

1,099

572

186

153

318

237

197

2,416

2,843

500

838

-63

186

153

3,760

4,307

5,074

1,172

1,416

3,084

146

182

215

1,318

1,598

3,299

2,442

2,709

1,775

®Lugo et al. (1975), Twilley (1982, 1985), and Twilley et al. (1986).
"Golley et al. (1962).

^Carter et al. (1973).

Source’. Twilley (1988)

of sediment accumulation or erosion, and by indirectly regenerating nutrients lost
from the root zone. The principal chemical conditions that affect primary productiv-
ity are soil water salinity and the concentration of major nutrients. High soil salinities,
which, in turn, are a function of the local hydrology and geomorphology, appear to be
the most important variable that influences the productivity of mangroves in a given
region. For example, one study of mangroves in Puerto Rico (Cintron et al., 1978)
found that tree height of the mangroves, as a measure of productivity, was inversely
related to soil salinity, according to the following relationship:

h=16.6-0.20Q (9.2)

where

h = tree height (m)

= soil salinity concentration (ppt)

In a similar analysis, J. W. Day et al. (1996) were able to demonstrate a relationship
between total litterfaii in an Avicmnia-domm.’Si.tcd basin mangrove forest in Mexico
and soil salinity as

L= 3.915 -0.039Q

(9.3)

Ecosystem Function 329

where

L = litterfall (g m ^ yr

Lovelock (2008) compared mangrove productivity from 11 forests around the
world and found very good correlations between soil respiration and measures of leaf
production and biomass. However, the highest below- ground carbon allocation per
unit litterfall was found in scrub mangrove forests. It seems that mangroves allocate
more growth below-ground during environmentally stressful conditions.

Nutrient availability has also been shown to have a major influence on man-
grove productivity. Feller et al. (2007) experimentally relieved nutrient deficiencies in
mangroves forests that were phosphorus limited (in Twin Cays, Belize) and nitrogen
limited (Indian River Lagoon, Florida). When nutrient deficiency was relieved, black
mangroves {A. ^erminans) at both forests responded with enhanced stem growth with
the greatest response coming from the N-limited forest. Nutrient enrichment was
found to influence more ecological processes in the P-hmited Twin Cays forest. The
authors concluded that eutrophication is more likely to shift nutrient limitations there
than the N-limited Indian River Lagoon.

Hurricane Effects

Hurricanes (and typhoons) and mangrove swamps could be described as having a tur-
bulent tropical love-hate relationship. Either by genetic design or chance, the time
required for the attainment of a steady state in mangrove forest in Florida is approxi-
mately the same as the average period between tropical hurricanes (approximately 20
to 24 years for Caribbean systems). This match suggests that mangroves may have
adapted or evolved to go through one life cycle, on average, between major tropical
storms. On average, a mangrove forest reaches maturity just as the next hurricane or
typhoon hits.

When Hurricane Andrew passed over south Florida in the late summer of 1992,
an opportunity existed for detailed studies of the immediate impact of hurricanes on
mangroves. Major damage to mangroves in the vicinity of the Everglades occurred due
to trunk snapping and uprooting rather than due to any storm surge (T. J. Smith et al.,
1994). Mortality was greatest for red mangroves in the 15- to 30-cm dbh age class
and over a wider band of 10 to 35 cm dbh for black and white mangrove. T. J. Smith
et al. (1994) made two other interesting observations:

1 . Gaps that developed in the mangrove forests due to lightning prior to the
hurricane were, after the hurricane, small green patches amid the gray matrix
of dead mangroves. Apparendy, the small-sized mangrove trees that were
spared in these patches would now serve as propagule regeneration sites for
the area of catastrophic disturbance, which was of much larger scale.
Small-scale disturbance nested in large-scale disturbance provides a positive
feedback for more rapid mangrove recovery.

330 Chapter 9 Mangrove Swamps

2. The loss of mangrove trees in a hurricane removes a major source of aeration
of mangrove soils, the trees themselves. As a result of the loss of this aeration,
soils in hurricane -impacted mangrove wetlands might become even more
reduced, producing even more toxic hydrogen sulfide that could preclude
mangrove regeneration for a number of years. Because of this negative
feedback, recovery of the mangrove forest to a system similar to that prior to
the disturbance is not assured.

Other research after Hurricane Andrew has looked at the vegetative response.
Monitoring forest plots inside and outside the eye -wall path of the hurricane between
1995 and 2005, Ward et al. (2006) found that turnover rates (mortality and recruit-
ment) between the two sections differed (dynamic turnover rates were greater inside
the eye-wall). However, both sections also exhibited a steady rate of forest turnover
dynamics, suggesting that ecological conditions and not structural conditions primar-
ily control productivity. In the case of a large climatic disturbance such as hurricanes,
the deposition of coarse woody debris may constitute a substantial nutrient flux.

Response to Rising Sea Levels

Sea-level rise has been documented around the world, and many coastal ecosystems
are expected to be adversely affected. The extent that sea-level rise may displace man-
grove forests is unclear, however, because these ecosystems are naturally adapted to
be resilient to disturbance (Krauss et al., 2014). Along! (2008) pointed out that man-
groves have several characteristics which make them resilient to disturbances that are
acute (e.g., hurricanes, tsunamis) or chronic (changing sea level). These characteristics
include: below-ground reservoirs of nutrients, rapid nutrient flux and decomposition,
complex and efficient biotic controls, and often rapid reconstruction following distur-
bance due to the self-design and simple architecture of these forests.

It appears that some mangroves forests may be more capable of acclimating to a
rising sea level than others. Many of the world’s mangrove forests are vulnerable to cli-
mate change, including those in the Caribbean and the Pacific Islands. Along! (2008)
predicted that the most susceptible mangrove forests will be slower- growing forests at
the thresholds of their habitat range, such as mangrove forests in arid regions where
mangroves grow more slowly due to higher salinities, lower humidity, and extreme
light conditions. Other susceptible mangroves will include forests in carbonate envi-
ronments where growth is slow and the input of terrestrial sediment is limited.

Based on current projections of climate change and sea-level rise. Along! (2008)
predicted a decline of 10 to 15 percent of the global mangrove area by 2100. Although
this decline is considerable, the author pointed out that the threat may be moot if
current mangrove deforestation trends continue worldwide.

Organic Material Storage and Export

Donato et al. (2011) made an interesting comparison that tropical mangrove are
among the most carbon-dense ecosystems in the world. They found that tropical

Ecosystem Function 331

1,400
-p 1,200
I’ 1,000

^ 800
O
%

O 600

E
0)

% 400

lij 200
0

Figure 9.7 Carbon storage in Indo-Pacific tropical mangrove forests compared to boreal,
temperate, and tropical forests. (After Donato et al., 2011)

mangroves had carbon storage of over 1,000 Mg ha”^ (10,000 g-C mT^) when soil
carbon below the 30 cm root zone is included, far more than the carbon storage in
tropical, temperate, or boreal forests (Fig. 9.7). They extrapolated, using a world cov-
erage of 140,000 km^ of mangroves, that mangrove wetlands store on the order of 4
to 20 petagrams (x 10^^ g) of carbon (Pg-C) globally.

Mangrove swamps are important exporters of organic material to the adjacent
estuary through the same ouWdlin£i dxscwsscd for coastal marshes in Chapter 8: “Tidal
Marshes. ” In one of the first studies on outwelling from mangroves, Heald ( 1971 ) esti-
mated that about 50 percent of the aboveground productivity of a mangrove swamp
in southwestern Florida was exported to the adjacent estuary as particulate organic
matter (POM). From 33 to 60 percent of the total POM in the estuary came from
Rhizophom (red mangrove) material. The production of organic matter was greater
in the summer (the wet season in Florida) than in any other season, although detrital
levels in the swamp waters were greatest from November through February, which
is the first four months of a typical seven month dry season. Thirty percent of the
yearly detrital export occurred during November. Heald also found that as the debris
decomposed, its protein content increased. The apparent cause of this enrichment,
also noted in salt marsh studies, is the increase of bacterial and fungal populations.

Since those early studies, an abundance of studies have been undertaken on the
outwelling from mangroves to adjacent estuaries. Almost all of the studies agree that
there is export of particulate organic carbon from mangroves to the adjacent estuary.
A summary of several studies on carbon export from mangrove wetlands suggested an
average of about 200 g-C rcT^ yi'^\ about double that exported from salt marshes
(Twilley, 1998). In a comparison of leaf litter production and organic export for
riverine, fringe, and basin mangrove systems (Fig. 9.8), riverine mangrove systems
exported a majority of their organic litter (94 percent, or 470 g-C m“^ yr“^ ), whereas
basin mangroves exported much less (21 percent, or 64 g-C rcT^ yr ^), leaving the

Boreal Temperate Tropical Mangrove

upland Indo-Pacific

332 Chapter 9 Mangrove Swamps

500

456

300

Riverine Fringe Basin

Mangrove Mangrove Mangrove

Figure 9.8 Organic carbon fluxes through mangrove swamps: infiows (litterfaii), export to
adjacent aquatic systems, and other iosses (decomposition and peat production). Width of
each pathway is proportional to flow (in g-C m~^ yr~^). (After Twilley et al., 1986)

leaf litter to decompose or accumulate as peat. The proportion of litterfaii production
that is exported and the total amount of litter that is exported increase as the tidal
influence increases.

Carbon Sequestration

A great deal of attention in the last decade has been paid to the sequestration or burial
of carbon by coastal wetlands, even to the point where that carbon is now referred
to as “blue carbon” (Mcleod et al., 2011). Carbon accumulation in mangrove soils is
a significant part of this coastal wetiand carbon accumulation. The most recent meta
estimates for carbon sequestration by mangrove swamps range from 160 ±40 g-C m“^
yr“^ (Breithaupt et al., 2012) to 226 ± 39 g-C yr^'^ (Mcleod et al., 2011). These
rates are 30 to 50 times the rate of carbon burial estimated for terrestrial uplands
forests (Mcleod et al., 2011). These mangrove estimates, when multiplied by area of
mangroves in the world described earlier, suggest that mangroves could be responsible
for 26 to 34 Tg-C/yr, a small fraction of the total carbon sequestered by the world’s
wetlands as described in Chapter 17: “Wetlands and Climate Change.” Breithaupt
et al. (2014) estimated a slightly lower rate of organic carbon sequestration by man-
groves of 123 ± 19 (std dev.) g-C mr^ yr“^ in the southern Florida Everglades. Because
this is an inorganic carbon-rich region, it would not be surprising if the total carbon
sequestration into the soil is significantly higher than this value.

Effects on the Estuary

The role of mangrove wetlands as both a habitat and a source of food for estuarine
fisheries is one of the most often-cited functions of these ecosystems. The fact that

Ecosystem Function 333

Figure 9.9 Detritus-based food web in south Florida estuary showing major contribution of
mangrove detritus to fisheries and the estuarine food chain. (After W. E. Odum and Heald,
1972)

organic carbon is exported from mangrove wetlands does not guarantee that it enters
estuarine food chains. Yet several independent studies have verified that mangrove
wetlands are important nursery areas and sources of food for sport and commercial
fisheries (Fig. 9.9 and 9.10).

Early studies by W. E. Odum (1970) and W. E. Odum and Heald (1972) estab-
lished that detrital export is important to sport and commercial fisheries (Fig. 9.9).
Through the examination of the stomach contents of more than 80 estuarine animals,
Odum found that mangrove detritus, particularly from Rhizophom, is the primary
food source in the estuary. Important consumers include the spiny lobster {PanuUrus
arpfus)^ pink shrimp {Penaeus duorarum)^ mullet (Mugil cephalus), tarpon {Mejjalops
atlanticm), snook {Centropomus undecimalis), and mangrove snapper {Lutjanus apo-
dus). The primary consumers also used the mangrove estuarine waters during their
early life stages as protection from predators and as a source of food.

334 Chapter 9 Mangrove Swamps

Eggs

Young

Spawning | Archosargusrhomboidalis
adults

Feb.-May | Haemulon plumieri

Pre-spawning adults
Aug.-Jan.

Subadults
Mar, -June

I Urolophusjamaicenss

Spawning adults
June-Sept. (seagra^s)
Juveniles |

Aug.-Dee. (mangroves)
Subadults •

Nov-May I

-To Sea

I Qchlasoma urophthalmus \

Spawning adults
June-Sept. (mangroves)
Juvenlles/sub-adults
Oct.-Feb. (seagrasses)
Prespawning adults
Feb.-May

Inflowing

fresh

water river

Figure 9.10 Life histories and habitat utilization of six fish species including
marine-estuarine spawners, estuarine spawners, and freshwater spawners in a mangrove
lagoon on the Gulf of Mexico, Mexico. (After Yanez-Arancibia et al., 1988)

In a series of studies in Terminos Lagoon in Mexico (Yariez-Arancibia et al., 1988,
1993), the use of mangrove forests by estuarine fish was clearly illustrated (Fig. 9.10).
There are clear connections among seasonal pulsing of mangrove detrital production,
adjacent planktonic and seagrass productivity, and fish movement and secondary pro-
ductivity. It is reasonable to extrapolate from this and similar studies that the removal
of mangrove wetlands would cause a significant decline in sport and commercial fish-
eries in adjacent open waters.

Organic carbon export from mangroves includes not only POM, which is most
often measured, but also dissolved organic matter, which has not received as much
attention as POM in export studies. Wafar et al. (1997) measured all of these fluxes
as well as those of particulate and dissolved nitrogen and phosphorus in a Rhizophora
swamp on the west coast of India. They made the following three conclusions:

1 . Mangrove production is important only for the carbon budget of the
adjacent estuary, not for the nitrogen or phosphorus budget.

2. The energy flux coming from the mangroves is more important for sustaining
microbial food chains than for sustaining particulate food chains.

References 335

3. The influence of the mangrove forest on the adjacent estuary relative to

phytoplankton production in the estuary is a function the size of the adjacent
estuary relative to the size of the mangrove forests.

Recommended Readings

Lee, S. Y., J. H. Primavera, F. Dahdouh-Guebas, K. McKee, J. O. Bosire, S. Cannicci,
K. Diele, F. Fromard, N. Koedam, C. Marchand, I. Mendelssohn, N. Mukherjee,
and S. Record. 2014. Ecological role and services of tropical mangrove ecosys-
tems: A reassessment. Global Ecolojjy and Biojjeojiraphy 23: 726-743.

Twilley, R. R, and J. W. Day. 2013. Mangrove Wetlands (pp. 165-202). In J. W.
Day, B. C. Crump, W. M. Kemp, and A. Yanez-Arancibia, eds.. Estuarine Ecolojjy.
Hoboken, NJ : Wiley-Blackwell.

References

Allen, J. A. 1998. Mangroves as alien species: The case of Hawaii. Global Ecology and
Biojjeopiraphy Letters 7: 61-71.

Alongi, D. M. 2008. Mangrove forests: resilience, protection from tsunamis, and
responses to global climate change. Estuarine, Coastal and Shelf Science 76: 1-13.
Bouillon, S., A. V. Borges, E. Castaheda-Moya, K. Diele, T. Dittmar, N. C. Duke, E.
Kristensen, S. Y. Lee, C. Marchand, J. J. Middelburg, V. H. Rivera-Monroy, T.
J. Smith 111, and R. R. Twilley. 2008. Mangrove production and carbon sinks: A
revision of global budget estimates. Global Bio^eochemistry Cycles 22:GB2013.
Breithaupt, J. L., J. M. Smoak, T. J. Smith, C. J. Sanders, and A. Hoare 2012, Organic
carbon burial rates in mangrove sediments: Strengthening the global budget.
Global Biojjeochem. Cycles 26: GB3011.

Breithaupt, J. L., J. M. Smoak, T. J. Smith, C. J. Sanders. 2014. Temporal variability of
carbon and nutrient burial, sediment accretion, and mass accumulation over the
past century in a carbonate platform mangrove forest of the Florida Everglades.
/. Geophysical Research: Biojjeosciences 119: 2032-2048.

Carter, M. R., L. A. Bums, T. R. Cavinder, K. R. Dugger, P. L. Fore, D. B. Hicks, H. L.
Revells, and T. W. Schmidt. 1973 . Ecosystem Analysis of the Bijj Cypress Swamp and
Estuaries, U.S. EPA 904/9-74[sim]]02, Region IV, Atlanta.

Cavanaugh, K. C., J. R. Kellner, A. J. Forde, D. S. Gruner, J. D. Parker, W. Rodriguez,
and 1. C. Feller. 2014. Poleward expansion of mangroves is a threshold response to
decreased frequency of extreme cold events. Froc. Nat. Acad. Sci. Ill: 723-727.
Chapman, V. J. 1976. Mangrove Vegetation, J. Cramer, Vaduz, Germany. 447 pp.
Chapman, V. J. 1977. Wet Coastal Ecosystems. Elsevier, Amsterdam, 428 pp.

Cintron, G., A. E. Lugo, D. J. Pool, and G. Morris. 1978. Mangroves of arid envi-
ronments in Puerto Rico and adjacent islands. Biotropica 10: 110-121.

Cintron, G., A. E. Lugo, and R. Martinez. 1985. Structural and functional properties
of mangrove forests. In W. G. D’Arcy and M. D. Corma, eds.. The Botany and
Natural History of Panama, IV Series: Monographs in Systematic Botany, vol. 10.
Missouri Botanical Garden, St. Louis, pp. 53-66.

336 Chapter 9 Mangrove Swamps

Cole, T. G., K. C. Ewel, and N. N. Devoe. 1999. Structure of mangrove trees and
forests in Micronesia. Forest Ecolo£iy and Mana£iement 117: 95-109.

Davis, J. H. 1940. The ecology and geologic role of mangroves in Florida. Publication
517. Carnegie Institution of Washington^ 303-412.

Day, J. W., Jr., C. Coronado-Molina, F. R. Vera-Herrera, R. Twilley, V. H.
Rivera-Monroy, H. Alvarez-Guillen, R. Day, and W. Gonner. 1996. A 7-year
record of above-ground net primary production in an southeastern Mexican
mangrove forest. Aquatic Botany 55: 39-60.

Dent, D. F. 1986. Acid Sulphate Soils: A Baseline for Research and Development . IFRl
Publication 39, Wageninggen, Netherlands.

Dent, D. F. 1992. Reclamation of acid sulphate soils. In R. Fal and B. A. Stewart,
eds. Soil Restoration, Advances in Soil Science ^vo\. 17. Springer-Verlag, New York,
pp. 79-122.

Donato, D. G., J. B. Kauffman, D. Murdiyarso, S. Kurnianto, M. Stidham, and M.
Kanninen. 2011. Mangroves among the most carbon-rich forests in the tropics.
Nature Geoscience Letters 4: 293-297.

Fgler, F. F. 1952. Southeast saline Everglades vegetation, Florida, and its manage-
ment. Vegetatio Acta Geobotica 3: 213-265.

Ewel, K. G., R. R. Twilley, and J. E. Ong. 1998. Different kinds of mangrove forests
provide different goods and services. Global Ecology and Biogeography Letters 7 :
83-94.

Ellison, A. M., E. J. Farnsworth, and R. E. Merkt. 1999. Origins of mangrove ecosys-
tems and the mangrove biodiversity anomaly. Global Ecology and Biogeography 8 :
95-115.

Feller, 1. G., G. E. Fovelock, and K. F. McKee. 2007. Nutrient addition differentially
affects ecological processes o? Avicenniagerminans'm nitrogen versus phosphorus
limited mangrove ecosystems. Ecosystems 10: 347-359.

Giri, G., E. Ochieng, F.F. Tieszen, Z. Zhu, m A. Singh, T. Foveland, J. Masek, and
N. Duke. 2011 Status and distribution of mangrove forests of the world using
earth observation satellite data. Global Ecology and Biogeography 23: 154-159.

Golley, F. B., H. T. Odum, and R. F. Wilson. 1962. The structure and metabolism of
a Puerto Rican red mangrove forest in May. Ecology 43: 9-19.

Heald E. J. 1971 . The Production of Organic Detritus in a South Florida Estuary. Uni-
versity of Miami Sea Grant Technical Bulletin No. 6, Goral Gables, FT, 110 pp.

Knight, J. M., P. E. R. Dale, R. J. K. Dunn, G. J. Broadbent, and G. J. Ecmckert. 2008.
Patterns of tidal flooding within a mangrove forest: Goombabah Fake, Southeast
Queensland. Estuarine, Coastal and Shelf Science 76:580-593.

Krauss, K. W., K. F. McKee, G. E. Fovelock, D. R. Gaboon, N. Saintilan, R. Reef, and
F. Ghen. 2014. How mangrove forests adjust to rising sea level. New Phytologist
202: 19-34.

Fee, S. Y, J. H. Primavera, F. Dahdouh-Guebas, K. McKee, J. O. Bosire, S. Gannicci,
K. Diele, F. Fromard, N. Koedam, G. Marchand, 1. Mendelssohn, N. Mukherjee,

References 337

and S. Record. 2014. Ecological role and services of tropical mangrove ecosys-
tems: A reassessment. Global Ecology and Bioj^eojjraphy 23: 726-743.

Lovelock, C. E. 2008. Soil respiration and belowground carbon allocation in man-
grove forests. Ecosystems 11: 342-354.

Lugo, A. E., G. Evink, M. M. Brinson, A. Broce, and J. C. Snedaker. 1975. Diurnal
rates of photosynthesis, respiration, and transpiration in mangrove forests in South
Florida. In F. B. Golley and E. Medina, eds. Tropical Ecolopfical Systems — Trends
in Terrestrial and Aquatic Research. Springer-Verlag, New York, pp. 335-350.

Lugo, A. E. 1980. Mangrove ecosystems: successional or steady state? Biotropica (sup-
plement) 12: 65-72.

Lugo, A. E. 1988 . The mangroves of Puerto Rico are in trouble. Acta Cientijical-. 124.

Lugo, A. E., and G. Patterson-Zucca. 1977. The impact of low temperature stress on
mangrove structure and growth. Tropical Ecology 18: 149-161.

Mcleod, E., G. L. Ghmura, S. Bouillon, R. Salm, M. Bjork, G. M. Duarte, G. E.
Lovelock, W. H. Schlesinger, and B. R Silliman. 2011. A blueprint for blue car-
bon: toward an improved understanding of the role of vegetated coastal habitats
in sequestering GO2. Frontiers of Ecology and the Environment 9: 552-560.

Odum, E. P. 1969. The strategy of ecosystem development. Science 164: 262-270.

Odum, W. E. 1970. Pathways of Enerp;y Flow in a South Florida Estuary. Ph.D. disser-
tation, University of Miami, Goral Gables, Florida, 62 pp.

Odum, W. E., and E. J. Heald. 1972. Trophic analyses of an estuarine mangrove
community. Bulletin of Marine Science 22: 671-738.

Odum, W. E., G. G. Mclvor, and T. J. Smith. 1982. The Ecology of the Mangroves
of South Florida: A Community Profile. U.S. Fish & Wildlife Service, Office of
Biological Services, Technical Report FWS/OBS 81-24, Washington, DG.

Patterson, S. G. 1986. Mangrove Community Boundary Interpretation and Detection
of Areal Changes in Marco Island, Florida: Application of Digital Image Processing
and Remote Sensing Techniques. Biological Services Report 86 (10), U.S. Fish and
Wildlife Service, Washington, DG.

Rabinowitz, D. 1978. Dispersal properties of mangrove propagules. Biotropica 10:
47-57.

Robertson, A. 1., and P. A. Daniel. 1989. The influence of crabs on litter processing
in high intertidal mangrove forests in tropical Australia. Oceologia 78: 191-198.

Schaeffer-Novelli, Y, G. Gintron-Molero, R. R. Adaime, and T. M. de Gamargo.
1990. Variability of mangrove ecosystems along the Brazilian coast. Estuaries 13:
204-218.

Smith, T. J., H-T. Ghan, G. G. Mclvor, and M. B. Robblee. 1989. Gomparisons of
seed predation in tropical, tidal forests on three continents. Ecology 70: 146-151.

Smith, T. J., K. G. Boto, S. D. Frusher, and R. L. Giddins. 1991. Keystone species and
mangrove forest dynamics: The influence of burrowing by crabs on soil nutrient
status and forest productivity. Estuarine, Coastal, and Shelf Science 33: 419-432.

Smith, T. J., M. B. Robblee, H. R. Wanless, and T. W. Doyle. 1994. Mangroves,
hurricanes, and lightning strikes. BioScience 44: 256-262.

338 Chapter 9 Mangrove Swamps

Stewart, G.R., and M. Popp. 1987. The ecophysiology of mangroves. In R. M. M.
Crawford, ed.. Plant Life in Aquatic and Amphibious Habitats^ Spec Publ. British
Ecological Society, vol. 5, pp. 333-345.

Thibodeau, F. R., and N. H. Nickerson. 1986. Differential oxidation of mangrove
substrate by Avicennia ^erminas and Rhizophora mangle. American Journal of
Botany 73: 512-516.

Thom, B. G. 1982. Mangrove ecology: A geomorphological perspective. In Clough,
B.F., ed.. Mangrove Ecosystems in Australia. Australian National University Press,
Canberra, pp. 3-17.

Twilley, R. R. 1982. Fitter dynamics and organic carbon exchange in black man-
grove {Avicennia germinans) basin forests in a Southwest Florida estuary. Ph.D.
dissertation. University of Florida, Gainesville.

Twilley, R. R. 1985. The exchange of organic carbon in basin mangrove forests in a
southwest Florida estuary. Estuarine, Coastal and Shelf Science 20: 543-557.

Twilley, R. R. 1998. Mangrove wetlands. In M. G. Messina and W. H. Conner,
eds.. Southern Forest Wetlands: Ecology and Management . Fewis Publishers, Boca
Raton, FT, pp. 445^73.

Twilley, R. R., A. E. Fugo, and C. Patterson-Zucca. 1986. Fitter production and
turnover in basin mangrove forests in southwest Florida. Ecology 67: 670-683.

Twilley, R. R., R. H. Chen, and T. Hargis. 1992. Carbon sinks in mangroves and their
implications to carbon budget of tropical coastal ecosystems. Water, Air, and Soil
Pollution 64: 265-288.

Twilley, R. R., and J. W. Day, Jr. 2013. Mangrove wetlands. In. J. W. Day, Jr., B. C.
Crump, W. M. Kemp, and A. Yanez-Arancibia, eds.. Estuarine Ecology, 2nd ed.
Wiley-Blackwell, Hoboken, NJ, pp. 165-202.

Wafer, S., A. G. Untawale, and M. Wafar. 1997. Fitter fall and energy flux in a man-
grove ecosystem. Estuarine, Coastal and Shelf Science 44: 111-124.

Ward, G. A., T. J. Smith III, K. R. T. Whalen, and T. W. Doyle. 2006. Regional pro-
cess in mangrove ecosystems: spatial scaling relationships, biomass, and turnmover
rates following catastrophic disturbance. Hydrobiologia 569: 517-527.

Wharton, C. H., H. T. Odum, K. Ewel, M. Duever, A. Fugo, R. Boyt, J.
Bartholomew, E. DeBellevue, S. Brown, M. Brown, and F. Duever. 1976.
Forested Wetlands of Florida — Their Management and Use. Center for Wedands,
University of Florida, Gainesville, 421 pp.

Yanez-Arancibia, A., A. F. Fara-Dominguez, J. F. Rojan-Galaviz, P. Sanchez-Gil, J. W.
Day, and C. J. Madden. 1988. Seasonal biomass and diversity of estuarine fishes
coupled with tropical habitat heterogeneity (southern Gulf of Mexico). Journal
of Fish Biology 33 (Suppl. A): 191-200.

Yanez-Arancibia, A., A. F. Fara-Dominguez, and J. W. Day. 1993. Interactions
between mangrove and seagrass habitats mediated by estuarine nekton assem-
blages: Coupling of primary and secondary production. Hydrobiologia 264:
1-12.

Freshwater marsh in Wisconsin

Chapter 10

Freshwater Marshes

Approximately 90 to 95 percent of the world’s wetlands are inland, or
nontidal. Inland freshwater marshes are perhaps the most diverse of the
wetland types discussed in this book and include the pothole marshes of the
north-central United States and south-central Canada, the Florida
Everglades, many wetland expanses in the Pantanal in Brazil, and the
floodplains of the Okavanpyo Delta in Botswana. Vepietation in freshwater
marshes is characterized by tall p/raminoids such as Typha and PhragmiteSj the
grasses Panicum and Cladiunij the sedges Scirpus^ Schoenoplectus^ Cyperus^
and CareXj broad-leaved monocots such Sagittaria spp., and floating aquatic
plants such ajNymphaea and Ndumbo. Some inland marshes, such as the
prairie pothole marshes, follow a cycle that includes dr oup/ht, refloodinp!, and
herbivory. In contrast to bopfs, mineralsoil-based inland marshes have hipfh-pH
substrates, hipfh available soil calcium, medium or hipfh loadinp; rates for
nutrients, often hipfh productivity, and hipfh soil microbial activity that leads to
rapid decomposition, recycling, and nitropfen fixation. Most of the primary
productivity is routed throupfh detrital pathways, but herbivory can be
seasonally important, particularly by muskrats and^eese. Inland marshes are
valuable as wildlife islands in the middle of agricultural landscapes and have
been tested extensively as sites for assimilating nutrients.

Most of the wetlands of the world are not located along the coasdines but are found
in interior regions. (These wedands are called “nonddal” in coastal regions to distin-
guish them from coastal wetlands.) We esdmate that there are about 5.5 million km^
of inland wetlands in the world (Table 10.1); in other words, they make up about
95 percent of all the world’s wetlands. About 415,000 km^ (about 95 percent) of the
total wedands in the conterminous United States are inland. This esdmate includes

341

342 Chapter 10 Freshwater Marshes

Table 10.1 Estimated area of inland wetlands in the world and North America (x 1,000 ha)

Peatlands

Freshwater Marshes

Freshwater Swamps

Total

World

350, 000'*

95,000*’

109,000"

554,000

North America

Conterminous U.S."'

3,700

9,600

28,200

41,500

Alaska'

51,800

17,000

68,800

Canada®

110,000

15,900

125,900

®Bridgham et al. (2001).

'’Average of several independent estimates.

"Matthews and Fung (1987).

"'Dahl (2006); freshwater marshes includes freshwater emergent and freshwater nonvegetated wetlands.
"All palustrine forested wetlands in Alaska and Canada are assumed to be peatlands.

'Hall etal. (1994).

*Zoltai (1988).

about 26,000 km^ of nonvegetated freshwater ponds. Including Alaska, there are 1.1
million km^ of inland wetlands in the United States.

It is difficult to put these inland wetlands into simple categories. Our simplified
scheme divides them into three groups: freshwater marshes (this chapter), tfeshwater
swamps (Chapter 11: “Freshwater Swamps and Riparian Ecosystems”), and boreal
peatlands (Chapter 12: “Peatlands”). These divisions roughly parallel the divisions
that persist in both the scientific literature and the specializations of wetland scientists.

Terminology for wedands, especially inland freshwater wetlands, can be confus-
ing and contradictory. In Europe, for example, the term reed swamp is often used to
describe one type of tfeshwater marsh dominated by Phr a£[mites spp., whereas in the
United States, the word swamp usually refers to a forested wetland. We consider Phrajj-
mites reed swamps to be marshes or marshhke, and they are covered in this chapter.
Although the use of classifying terms connotes clear boundaries between different
wetland types, in reality they form a continuum. The extremes of freshwater marshes
are clearly different, but at the boundaries between two wetland types (e.g., marsh
and bog), the distinction is not always clear. Marshes (and reed swamps) have mineral
soils rather than peat soils. American terminology has developed without much regard
to whether the system is peat forming. The fact is that most freshwater marshes and
swamps, regardless of where they are located and regardless of their geological origins,
accumulate some peat.

Freshwater marshes includes a diverse group of wetlands characterized by
( 1 ) emergent soft-stemmed aquatic plants, such as cattails, arrowheads, pickerelweed,
reeds, and several other species of grasses and sedges; (2) a shallow-water regime; and
(3) generally shallow to nonexistent peat deposits. There are few accurate measures
of how many freshwater marshes there are in the world for several reasons. First,
they are often ephemeral or convert to other types of wetlands, such as unvegetated
fiats or forested wetlands, over a relatively short period of time. Second, they can be
confused with and miscounted with peatlands, particularly fens. Third, freshwater
marshes have such a wide number of possible dominant vegetation covers and water

Hydrology 343

depths (from saturated soils to 1 m of water depth) that their classification and
inventory is very difficult. We estimate that there are about 950,000 km^ million ha
of freshwater marshes in the world (Table 10.1), less than 20 percent of the total
amount of wetlands in the world. The Okavango Delta in Botswana, the Danube
and Volga Deltas in Eastern Europe, the Mesopotamian Marshlands in Iraq, the
Everglades in Florida, the Prairie Pothole region of the United States and Canada,
and the Pantanal in South America are all examples of regions with extensive expanses
of freshwater marshes. Freshwater marshes are estimated to cover about 96,000 km^
in the coterminous United States (Table 10.1).

Hydrology

As with any other wedand, the flooding regime, or hydroperiod, of freshwater marshes
determines their ecological character. The critical factors that determine the character
of these wedands are the presence of excess water and sources of water other than
direct precipitadon. The hydroperiods of several freshwater marsh systems were illus-
trated in Chapter 4: “Wedand Hydrology”. Along seacoasts, water levels tend to be
stable over the long term because of the influence of the ocean. Water levels in inland
marshes, in contrast, are much more controlled by the balance between precipitadon
and evapotranspiradon, especially for marshes in small watersheds that are affected
by restricted throughflow. Water levels of marshes, such as those found along the
Laurendan Great Lakes, are generally stable but are influenced by the year-to-year
variability of lake levels and by whether the wetland is diked or open to the lake.
Many marshes, such as wet meadows, sedge meadows, vernal pools, and even prairie
potholes, dry down seasonally, but the plant species found there reflect the hydric
condidons that exist during most of the year. The seasonality of these marshes is
due to the fact that they are fed primarily by runoff and precipitadon. Some marshes
intercept groundwater supplies. Their water levels, therefore, reflect the local water
table, and the hydroperiod is less erradc and seasonal. These types of marshes, such
as those found in the prairie pothole region of North America, can be either recharge
or discharge wetlands. Other marshes collect surface water and entrained nutrients
from watersheds that are large enough to maintain hydric condidons most of the
rime. For example, overflowing lakes and rivers supply water and nutrients to adja-
cent riparian or littoral marshes. Because river and stream discharge, lake levels, and
precipitadon are often notoriously variable due to weather shifts from year to year, the
water regime of most inland marshes also varies in a way that is predictable only in a
statistical sense.

Even in the same region, water levels can respond differently to shifts m the
balance between precipitation and evapotranspiradon yearly (Fig. 10.1). Terms such
as ephemeral, temporary, seasonally semipermanent, and permanent can be used to
describe freshwater marshes. In addition, marshes can move through several of these
classes over the span of a few years. Thus, a marsh that would ordinarily be consid-
ered permanent might be in a “drawdown” phase that gives the appearance of an
ephemeral marsh.

344 Chapter 10 Freshwater Marshes

Q -0
-8

560.0

559.0

558.0

1 — I — I — ^ — I — I — I — I — ^ — I — I — ^ — I — r

J ^ I I I I I I I *'1 I I ^ I I L

^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^

Wetland PI

^ \ . J

\ A

1 1 ^ 1 1 1 1 1 1 1 1 ^ 1 ! ^ 1 1

557.0

556.0

555.0

wetland

ground

surface

wetland

ground

surface

544.0

Wetland P11

A -

543.0

542.0

1 1

. \ \

\ ^

\ ^
^ A

^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ L

80 82 84 86 88 90 92 94

Year

wetland
J^ground
96 surface

Figure 10.1 Water-level patterns in three wetlands in the prairie pothole region of North
America, illustrating the uneven effect that climate has on supposedly similar wetlands in
the same region, probably due to nonuniform groundwater effects on the wetlands. PDSI is
the Palmer Drought Severity Index and is a relative measure of climatic “wetness.” Its value
decreases during drought conditions. (After LaBaugh et al., 1996)

Biogeochemistry

The water and soil chemistry of freshwater marshes is dominated by a combination
of mineral rather peat soils, overlain with autochthonous inputs of organic matter
from the productivity of the vegetation. Given these conditions, there is still a
wide range of chemical possibilities for the water and soil in freshwater marshes
(Fig. 10.2). Conductivity as a measure of general salinity can range from less than

South Dakota

Minnesota

Iowa

Nebraska

_j I I I

10 100 1,000 10,000 100,000
Specific conductivity, )iS/cm@25°C

1,000,000

Figure 10.2 Fi'equency distribution of salinity, as measured by specific conductance, for
prairie lakes and wetlands in the United States and Canada as measured by several investiga-
tors. Concentrations shown for Nebraska are typical of inland salt marshes. (After LaBaugh,
1989)

345

346 Chapter 10 Freshwater Marshes

100 microSiemens per cm (pS cm“^) in soft-water freshwater marshes dominated by
rainfall to over 300,000 pS cm”^ in “inland salt marshes” dominated by saline seeps in
semiarid climates. Differences are related to the magnitude of dissolved salts, nutrients,
and other chemical inputs and to the relative importance of groundwater and surface
water inflow. Inland marshes are generally minerotrophic in contrast to bogs. That
is, the inflowing water has higher amounts of dissolved materials, including nutrients,
resulting from the presence of dissolved cations in streams, rivers, and groundwater
compared to bogs that are fed simply by rainfall. The organic substrate of freshwater
marshes, while shallow compared to bogs and fens, is saturated with bases, and as a
result the pH is close to neutral. Because nutrients are usually plentiful, productivity is
higher in freshwater marshes than it is in bogs, bacteria are active in nitrogen fixation
and litter decomposition, and turnover rates are high. The accumulation of organic
matter that does occur results from high production rates, not from the inhibition of
decomposition by low pH (as occurs in bogs).

Wetland hydrology has a tremendous influence on biogeochemistry and, coupled
with a diverse geomorphology, can provide considerable spatial heterogeneity within
the wetland. On the St. Louis River system in Minnesota and Wisconsin, Johnston
et al. (2001) found higher variability in available nutrients within wetlands than
between two wetlands with very different soil conditions (silty versus clayey). The
significant intra-wetland variability was attributed to the variety of geomorphic
features (levees, backwater zones) that provided very different hydrologic conditions
within both wedands.

Nutrient concentrations reported for sediments in inland freshwater marshes vary
widely, depending on the substrate, parent material, open or closed nature of the basin,
connection with groundwater, and even nutrient uptake by plants. Ion concentrations
of freshwater marshes are high, and water is generally in a pH range of 6 to 9. Organic
matter can vary from a very high content (75 percent), as can be found in freshwa-
ter marshes in coastal Louisiana, to a low content (10-30 percent) in marshes fed by
inorganic sediments from agricultural watersheds or open to organic export. Concen-
trations of total (as distinguished from available) nutrients are reflections of the kinds
of sediments in the marsh. Mineral sediments are often associated with high phos-
phorus content, for example, whereas total nitrogen is closely correlated to organic
content. Dissolved inorganic nitrogen and phosphorus — the elements that most often
limit plant growth — often vary seasonally from very low concentrations in the summer,
when plants take them up as rapidly as they become available, to high concentrations
in the winter, when plants are dormant but mineralization continues in the soil.

In many parts of the world where arid climates persist, inland marshes can be
saline rather than freshwater. These marshes, then, have characteristics of both coastal
salt marshes (because of the salinity) and inland marshes (because they are not tidal).
Good examples of these kinds of marshes are the fringe marshes around the Great
Salt Lake in Utah, the Salton Sea in Galifbrnia, and the Gaspian Sea in Eastern
Europe. There are also inland “salt marshes,” some as large as 15 to 20 ha, still

Transition to Uplands

Emergent Plant Zone

Open Water

Hydrology Peiiodtcally
flooded

Waterlogged or
shallow water

Shallow arxl Continuous standing
continuously wet water > 30 cm

Plants lowland grasses. sedges

woody plants, e g. . C^rax spp.
cottonwood Rypulus, arrowhead
willow Sagittaria sf)p.

cattail
Typha spp
bulrush
Sotrpus

Schoenopfectus

water lilies
Nymphaaa spp
pondweeds
R>tamogeton spp.
bladderwort
Utnculahaspp.

(a)

Emergent Plant

Euhydrophyte

Floating Island

Figure 10.3 Cross sections of vegetation through freshwater marshes, indicating plant
zones according to water depth and typical plants found in each zone for (a) temperate-zone
midwestern North America and (b) sub-Saharan Africa. Note floating marsh islands in African
wetlands, ((b) After Denny, 1993)

347

348 Chapter 10 Freshwater Marshes

found northwest of Lincoln, Nebraska, along Salt Creek and its tributaries. These
marshes are the result of saline seeps from deep groundwater and the consistently
high evapotranspiration/precipitation ratio of the region.

Vegetation

The vegetation of fresh inland marshes has been detailed in many studies. The dom-
inant species vary from place to place, but the number of genera common to all
locations in the temperate zone is quite remarkable. Common species include the
graminoids Phra£imites australis {= P. communis', reed grass), Typhasp'p. (cattail), Spar-
jjanium eurycarpum (bur reed). Zizania aquatica (= Z. palustris-, wild rice), Panicum
hemitomon, Cladium jamaicmsc, sedges Carexsp'p., Schoenoplectus tabernaemontani
(= Scripus validus-, bulrush), Scirpus fluviatilis (river bulrush), and Eleocharis spp.
(spike rush). In addition, broad-leaved monocotyledons such as Pontederia cordata
(pickerelweed) and Sagittaria spp. (arrowhead) are frequently found in freshwater
marshes. Herbaceous dicotyledons are represented by a number of species, typical
examples of which are Ambrosiascpp. (ragweed) and Polygonum spp. (smartweed). Fre-
quently represented also are such ferns as Osmunda rcgalis {joyA fern) and Thdypteris
palustris (marsh fern), and the horsetail, Equisetum spp. One of the most productive
species in the world is the tropical sedge Cyperus papyrus, which flourishes in marshes
and on floating mats in southern and eastern Africa.

Marsh Vegetation Zonation

These typical plant species do not occur randomly mixed together in marshes. Each has
its preferred habitat. Different species often occur in rough zones on slight gradients,
especially flooding gradients. Figure 10.3a illustrates the typical distribution of species
along an elevation gradient in a midwestern North American freshwater marsh. Sedges
(e.g., Carexspp., Scirpusspp.), rushes (Juncusspp.), and arrowheads {Sagittaria spp .)
typically occupy the shallowly flooded edge of a pothole. Two species of cattail ( Typha
latifolia and T. angustifolia) are common. The narrow-leaved species (T. angustifo-
lia) is more flood tolerant than the broad-leaved cattail ( T. latifolia) and may grow
in water up to 1 m deep. The deepest zone of emergent plants is typically vegetated
with hardstem bulrush {Scirpus acutus) and softstem bulrush {Schoenoplectus taber-
naemontani). Beyond these emergents, floating-leaved and submersed vegetation will
grow, the latter to depths dictated by light penetration. Typical floating-leaved aquatic
hydrophytes include rhizomatous plants, such as water lilies ( Nymphaea tuberosa or
N. odorata), water lotus {Nelumbo lutea), and spatterdock {Nuphar advena), and
stoloniferous plants, such as water shield {Brasenia schreberi) and smartweed {Poly-
gonum spp.). Submersed hydrophytes include coontail {Ceratophyllum demersum),
water millfoil {Myriophyllum spp.), pondweed {Potamogeton spp.), wild celery ( Vallis-
neria americana), naiad {Naj as spp.), bladderwort ( Utricularia spp.), and waterweed
{Elodea canadensis).

A unique structural feature of prairie pothole marshes is the 5- to 20-year cycle
of dry marsh, regenerating marsh, degenerating marsh, and lake that is related to

Vegetation 349

periodic droughts. During drought years, standing water disappears. Buried seeds in
the exposed mud flats germinate to grow a cover of annuals {Bidens, Polyjjonum, Cype-
rus, Rumex) and perennials {Typha, Scirpm, Spar^anium, Sa^ittaria). When rainfall
returns to normal, the mud flats are inundated. Annuals disappear, leaving only the
perennial emergent species. Submersed species (Rotamogeton^ Najas, Ceratophyllum,
Myriophyllum, Cham) also reappear. For the next year or more, during the regener-
ating stage, the emergent population increases in vigor and density. After a few years,
however, these populations begin to decline. The reasons are poorly understood, but
often muskrat populations explode in response to the vigorous vegetation growth.
Their nest and trail building can decimate a marsh. Whatever the reason, in the flnal
stage of the cycle, there is little emergent marsh; most of the area reverts to an open
shallow lake or pond, setting the stage for the next drought cycle. Wildlife use of these
wetlands follows the same cycle. The most intense use occurs when there is good inter-
spersion of small ponds with submersed vegetation and emergent marshes with stands
diverse in height, density, and potential food.

Figure 10.3b shows the plant zonation of a freshwater marsh/httoral zone in
sub-Saharan Africa. The shallow flooded emergent zone is dominated by Typha,
Phrapimites, and Cyperus papyrus. Typha taxonomy in Africa has been somewhat
confused, but it now appears that there are two distinct species — T. domingmsis
Pers. sensu lato in tropical and warm-temperate climates and T. capmsis Rohrb. in
more temperate climates of northern and southern Africa. Cyperus papyrus grows
in the emergent zone but also develops floating islands when it breaks free from
the shoreline. The euhydrophyte (true water plant) zone shown in Figure 10.3b
refers to the zone with rooted, floating-leaved, and submerged macrophytes (Denny,
1985). In Africa, examples of other plants in this zone are Chara^ Pontinalis.,
T^ymphaea.,CeratophyUum.,VaUisneria., Potamogeton, and Paspalidium. Thus, despite
these differences and vastly different climates, freshwater marshes around the world
share some common species and many common genera and are functionally much
the same.

Inland Salt Marshes

Where evapotranspiration exceeds precipitation and/or saline groundwater
seeps occur, inland salt marshes are often found. For example, the Nebraska
salt marshes located in eastern Nebraska near Lincoln (Fig. 10.4) support
many plant genera familiar to coastal salt marsh ecologists. The most saline
parts of these marshes are dominated by salt-tolerant macrophytes such as
saltwort {Salicornia rubra), sea blight {Suaeda depressa), and inland saitgrass
(Distichlis spicata), whereas the open ponds and their fringes are dominated
by plants such as sago pondweed (Potamogeton pectinatus), wigeon grass
(Ruppia maritima), prairie bulrush [Scirpus maritimus var. paludosus), and even
cattails {Typha angustifolia and T. latifolia). In California, the major species in
brackish marshes include pickleweed {Salicornia virginica) and alkali bulrush
{Scirpus robustus).

350 Chapter 10 Freshwater Marshes

Creek

Channel

Transition

Zone

Aquatic

Community

:sallmarsh aster
■Aster subulatjs
rwestem ironweed
: \tfrmnia
fasaculata
: Plains bluegrass
■Fbaanda

spearscale
Atnplex subspicata
foxtail barley
Hordeum jubatum
marsh elder
tvaanrHia
saltmarsh aster
Aster subulatus

'salt crest

saltwort Salicornia rubra
sea blite Suaedadepressa
inland saitgrass Dikichtis
spicata

sago porrdweed
Fbtamogeton
pectirratus
prairie bulrush
Sarpusrrrantimus
var. paludosus

Figure 10.4 Cross section through an iniand sait marsh in Nebraska, indicating piant
zones and typicai piants found in each zone. (After Farrar and Gersib, 1991)

Vegetation Seed Banks and Diversity

Seed banks and fluctuating water levels interact in complicated ways to produce veg-
etation communities in freshwater marshes (see Chapter 7: “Wetland Vegetation and
Succession”). As a general rule, seed germination is maximized under shallow water
or damp soil conditions, after which many perennials can reproduce vegetatively into
deeper water. For example, fluctuating water levels along the Laurentian Great Lakes
allowed greater diversity of plant types and species in the coastal marshes, and these
marshes sometimes have a density of buried seeds an order of magnitude greater than
that of inland prairie marshes (Keddy, 2010).

An Experiment in Hydroperiods, Seed Banks, and Marsh
Vegetation Diversity

The importance of the timing of flooding and drying was illustrated in an
experiment in Ohio involving uniform seed banks subjected to several hydrope-
riods. Although plant density and above-ground biomass were not affected by
the different hydroperiods, species composition, diversity, and richness were
affected (Fig. 10.5). Highest richness and diversity occurred in continuously
moist soils. Flooding followed by a drawdown to moist soils, as is a typical
hydroperiod in the midwestern United States, encouraged obligate wetland

Vegetation 351

species, whereas moist soil conditions followed by flooding encouraged the
growth of fewer wetland species and more annual species.

Hydroperiod Vegetation response

continuously flooded

May ’ Jun ' Jul ' Aug

lowest wetland plant diversity
high density of obligate wetland plants
perennials with no annuals

mid-range species diversity

high density of obligate wetland plants

perennials dominated; almost no annuals

mid-range species diversity

low density of obligate wetland plants

mid-range density of annuals

highest species diversity

low density of obligate wetland plants

highest density of annuals

Figure 10.5 Experimental results of four different growing season hydroperiods on a
common seed bank of annual and perennial freshwater marsh plants in central Ohio.
Continuously deep water favored low density of obligate wetland perennials; flooded
then moist, typical of natural hydroperiod for the midwestern United States, favored
perennial-dominated and diverse wetland plants; moist then flooded, typical of some
managed marshes around the Laurentian Great Lakes, favored fewer obligate wetland
plants and higher density of annuals; continuously flooded soil had highest diversity
but fewest obligate wetland plants and highest density of annuals. (From Johnson,
1998)

352 Chapter 10 Freshwater Marshes

Standing crop, g/0.25m^

Figure 10.6 Species richness versus vegetation biomass in 0.25-m^ quadrants from three
wetland areas in Ontario, Quebec, and Nova Scotia. (After Moore et al., 1989)

The particular species found in freshwater wetlands are also determined by many
other environmental factors. Nutrient availability determines, to a large degree,
whether a wetland site will support mosses or angiosperms (i.e., whether it is a
bog or a marsh) and what the species diversity will be. For freshwater marshes, it
is not obvious that highly fertile wetlands are highly diverse. In fact, most studies
of freshwater marsh plant diversity published in the literature suggest the opposite
conclusion. As an example, Moore et al. (1989) contrasted several fertile and infertile
sites (as measured by the plant standing crop) in eastern Ontario and found the
greatest species richness in marshes that had peak biomass between 60 and 400 g/m^
and much less plant richness at higher plant standing crop (>600 g/m^) (Fig. 10.6).
They also found rare species only at the infertile sites, suggesting that the conservation
of infertile wetlands should be part of overall wetland management strategies. Keddy
(2010) pointed out that rare species and high marsh diversity are now routinely
found in “peripheral” habitats, as shown in his centrifigual model for herbaceous
plants as described in Chapter 7 and shown for a Typha core habitat in Figure
7.10b. We agree with his conclusions for herbaceous plant wetlands (freshwater
marshes) that:

■ Peripheral habitats contain more biological diversity and higher number of
rare species.

■ The core habitat is dominated by a few species (e.g., Typha, Phrapfinites,
Scirpus, Papyrus).

■ Any factor that increases fertility or decreases disturbance will force peripheral
habitats into the usually undiverse core habitat.

Vegetation 353

Because many inland marshes are potholes that collect water that leaves only
by evaporation, salts may become concentrated during periods of low precipitation,
adversely affecting the growth of salt-intolerant species. In a review of 90 emergent
wetlands along the U.S. Great Lakes, Johnston et al. (2007) revealed that plant
forms were often indicators of wetland soil types. They found that submerged aquatic
vegetation tended to indicate silty soils, free-floating plants indicated clay soils, and
graminoids indicated sandy soils.

Invasive Species

Nonnative plant species are often a part of the vegetation of freshwater marshes, partic-
ularly in areas that have been disturbed. It has been hypothesized that tropical regions
are more susceptible to invasion than temperate regions because invading plants grow
much more rapidly and are more noticeable in the tropics than in temperate lati-
tudes. Plants such as Eichhornia crassipes {water hyacinth), Salvinia molesta (salvinia),
and Alternanthera philoxeroidses (alligator weed) have invaded tropical and subtrop-
ical regions of the world. E. crassipes can double the area that it covers in two weeks
and has choked many waterways that have received high nutrient loads for almost a
century. Although there are many theories about alien aquatic plants, there is some
validity to the concept that disturbed ecosystems are most susceptible to biologi-
cal invasions. Werner and Zedler (2002) found that sediment accumulation within
Wisconsin sedge meadows reduced tussock microtopography, promoted invasion by
Typha spp. or Phalaris arundinacea (reed canary grass), and reduced wetland species
richness.

In the freshwater marshes of the St. Lawrence and Hudson River valleys and in
the Great Lakes region of North America, Lythrum salicaria (purple loosestrife), a
tall purple-flowered emergent hydrophyte, spread at an alarming rate in the twenti-
eth century, causing much concern to those who manage these marshes for wildlife
(Stuckey, 1980; Balogh and Bookhout, 1989). The plant is aggressive in displacing
native grasses, sedges, rushes, and even Typha spp. Many freshwater marsh managers
have implemented programs designed to control purple loosestrife by chemical and
mechanical means. Other aquatic aliens, such as the submersed Hydrilla verticillata,
a plant native to Africa, Asia, and Australia, and Myriophyllum spicatum^ have invaded
open, shallow-water marshes in the United States (Steward, 1990; Galatowitsch et ah,
1999) but rarely compete well with emergent vegetation.

Phrapimites australis (Gav.) Trin. Ex. Steud., common reed, is considered an inva-
sive species in eastern North America, particularly in along the Atlantic coastline and
around the Laurentian Great Lakes, even though the plant has been in North Amer-
ica for more than 3,000 years, because of its aggressive expansion through brackish
(salinity <5 ppt) and freshwater marshes, especially in the past 50 years. Its spread
is attributed to increased disturbances, spread of more aggressive varieties from other
parts of the world, including Europe, and changes in hydrology and salinity patterns in
coastal estuarine systems (Philipp and Field, 2005). Saltonstall (2002) confirmed the
existence of a total of 27 haplotypes of which II (types A-H, S, Z, AA) are native to

354 Chapter 10 Freshwater Marshes

North America. Within North America, Types AA, F, Z and S are known historically
from the Northeast; types E, G, and H from the Midwest; and types A to D from
the South and Intermountain West. Two haplotypes, I and M, show worldwide dis-
tribution, with M the most common type in North America, Europe, and Asia. A new
subspecies, Phra^mites australis suhsp . americanus, has been identified and has been
shown to be distinctly different from the introduced and Gulf Goast lineages of P.
australis (Saltonstall et ah, 2004). This creates a difficult problem in wetland manage-
ment and restoration, because wetland managers must be able to distinguish between
the invasive “bad” Phra^mites and the native “good” ones. Major resources are used
to control this plant from spreading in eastern North America with techniques such
as burning and herbicide application. There is irony in the fact that the plant is taking
over North American wetlands while reed dieback of the same species in Europe is the
main concern for this species there.

Consumers

Perhaps one reason that small marshes of the prairie region and the western high plains
harbor such a rich diversity of organisms and wildlife is that they are often natural
islands in a sea of farmland. Gultivated land does not provide a diversity of either
food or shelter, and many animals must retreat to the marshes, which have become
their only natural habitats. In cases where flow from watersheds is seasonal, freshwater
marshes can serve as biological and hydrologic “oases” during low-flow and drought
conditions.

Invertebrates

Invertebrates, similar to amphibians, are the links between plants and their detritus, on
one hand, and animals such as fish, ducks and other birds, and even several mammals,
on the other. Insects make up much of the invertebrate taxa in freshwater marshes, and
their composition is often dictated by wetland hydrology and vegetation. Temporary
pools tend to be diverse with beetle and midge communities. Insect communities are
often productive because of the alternating wet-dry conditions, and many communi-
ties are regulated by biotic interactions. As marshes become more perennial, vegetation
for habitat structure and as a decaying substrate for becomes important. In open-water
sections, certain benthos and nektonic insects may also be important.

The most conspicuous invertebrates are the true flies (Diptera), which often make
one’s life miserable in the marsh. These include midges, mosquitoes, and crane flies.
However, in the larval stage, many of the insects are benthic. Midge larvae, which
are called bloodworms because of their rich red color, “are found submerged in bot-
tom soils and organic debris, serving as food for fish, frogs, and diving birds. When
pupae surface and emerge as adults, they are exploited as well by surface-feeding birds
and fish” (Weller, 1994). Odonata, represented by dragonflies and damselflies, are a
notable feature of freshwater marshes; their very presence generally indicates good
water quality. Grustaceans such as crayfish and mollusks such as snails can be common

Consumers 355

in some freshwater marshes. The former are food for large fish and mammals alike,
whereas the latter are often found grazing on mats of filamentous algae.

Temporal cycles and spatial patterns of invertebrate species and concentrations
reflect the natural seasonal cycle of insect growth and emergence superimposed on the
vegetation cycles. McLaughlin and Harris (1990) investigated insect emergence from
diked and undiked marshes along Lake Michigan and found more insects, more insect
biomass, and a greater number of taxa in diked marshes and the greatest numbers and
biomass in the sparsely vegetated zones of the wetlands rather than in open water or
dense vegetation. Kulesza and Holomuzki (2006) examined growth and survival of
the detritivorous amphipod HyaleUa azteca from a Lake Erie marsh. They compared
its use of Typha an^ustifolia and Phra^mites australis as substrate and found that both
plants supported adequate fungi growth, and the amphipods performed equally well.

Amphibians

Amphibians are an important group of organisms in freshwater marshes, often serving
as the link between insect populations and wading birds, mink, raccoons, and some
fish in complex food webs. Larval tadpoles, which can be quite abundant in some
freshwater marshes, eat small plants and animals and are, in turn, eaten by large fish
and wading birds. The adult frogs feast on emerging insects. Even terrestrial toads use
freshwater marshes as mating and breeding grounds in the spring. There has been con-
cern about declining amphibian populations; one of the causes that has been suggested
has been the loss of wetland habitat. Richter and Azous (1995) investigated the rela-
tionships between amphibian richness and variables such as wetland size, vegetation
type, presence of competitors and predators, hydrologic characteristics, hydroperiod
fluctuations, and land use. The variables that explained the highest correlation with
amphibian richness were water-level fluctuations and percentage of the watershed that
was urbanized. These data do not explain the exact cause of the loss of amphibians, but
urban pollution and stream and hydrological modifications appear to be likely causes.

Porej (2004) compared several created and restored wedands in central Ohio and
found that the presence of a shallow-sloped littoral zone, the absence of fish (often
caused by flooding limited to seasonal patterns in the wedand), and a high rado of
edge to area of wetlands (optimized where there are many small basins rather than
one large basin of the same area) are among the key physical and biological features
that support a diversity of amphibians. American toads [Bufo americanus), northern
leopard frogs {Rana pipiens), western chorus frogs {Pseudacris triseriata), gray tree
frogs {Hyla versicolor) and smallmouth salamanders {Ambystoma texanum) were posi-
dvely correlated with the presence of shallow littoral zones in these freshwater marshes
and ponds. Porej (2004) also found higher salamander richness in forested wedands
compared to freshwater marshes (natural or created) while frogs and toads had sim-
ilar richness in forested wedands and marshes that was lower than the richness in
newly created marshes (Table 10.2). He found a strong associadon between the pres-
ence of forest cover within 200 m of freshwater wedands and amphibian diversity in
agricultural landscapes, parricularly for spotted salamanders {Ambystoma maculatum).

356 Chapter 10 Freshwater Marshes

Table 10.2 Occurrence (percentage of wetlands occupied) of pond-breeding amphibians in
54 naturai (emergent and forested) and 42 created wetiands iocated in the Till Plains and
Giaciated Piateau ecoregions of centrai Ohio

Species

Natural

Emergent

Wetlands

Natural

Forested

Wetlands

Created

Wetlands

Bufo americanus/Bufo fow/eri(American/Fowler's toad)

Rana clamitans melanota (Green frog)

R. pipiens (Northern leopard frog)

R. catesbeiana (American bullfrog)

R. sylvatica (Wood frog)

Pseudacris crucifer (Spring peeper)

R triseriata (Western chorus frog)

Hyla versicolor (Gray treefrog)

Acris crepitans blanchardii (Blanchard's cricket frog)

Frogs and toads (ave ± st error), species per wetland

3.2 ± 0.3

3.0 ± 0.3

3.9 ± 0.3

Ambystoma tigrinum (Tiger salamander)

A. maculatum (Spotted salamander)

A. texanum (Smallmouth salamander)

A. jeffersonianum complex (Jefferson's salamander complex)

A. opacum (Marbled salamander)

Notophthalamus viridescens v. (Red-spotted newt)

Salamanders (ave + st error), species per wetland

1.0 ± 0.3

2.4 ± 0.2

0.3 ± 0.1

Total amphibians, species per wetiand

4.2 ± 0.5

5.4 ± 0.3

4.2 ± 0.3

Source’. Porej (2004)

Jefferson’s salamander complex {A. jeffersonicmum complex), smallmouth salaman-
ders (A. texanum), and wood frogs {Rana sylvatica).

Fish

One of the most difficult issues about which to generalize is whether freshwater
marshes support much fish life or indeed if they should. As a general rule, the deeper
the water in the marsh and the more open the system is to large rivers or lakes, the
more variety and abundance of fish that can be supported. The positive aspect of
freshwater marshes as habitats and nurseries for fish was investigated by Derksen
(1989) for a large Manitoba marsh complexand by Stephenson (1990) for Great Lakes
marshes. Derksen (1989) found extensive use of the marshes by northern pike {Esox
lucius) with emigration from the marsh occurring primarily in the autumn. Stephen-
son (1990) found a total of 36 species of fish in marshes connected to Lake Ontario,
including spawning adults of 23 species and the young-of-the-year of 31 species, indi-
cating the importance of these marshes for fish reproduction in the lake. Eighty-nine
percent of the species encountered were using the marshes for reproduction.

Common carp ( Cyprinus carpio) are able to withstand the dramatic seasonal and
did fluctuations of water temperature and dissolved oxygen typical of shallow marshes

Consumers 357

and are thus abundant in many inland wetiands. They affect marsh vegetation by direct
grazing, uprooting vegetation while searching for food, and causing severe turbidity
in the water column. For these reasons, carp are not considered desirable by many
freshwater marsh managers.

Mammals

A number of mammals inhabit inland marshes. The most noticed is probably the
muskrat {Ondatra zibethicus). This herbivore reproduces rapidly and can attain pop-
ulation densities that decimate the marsh, causing major changes in its character. Like
plants, each mammalian species has preferred habitats. For example, muskrats are
found in the most aquatic areas, the water vole (e.g., Microtus richardsoni in North
America) in overlapping but higher elevations, and other voles in the relatively ter-
restrial parts of the reed marsh. Most of the mammals are herbivorous. The effects
that beavers ( Castor canadensis) have on hydrology of wetlands and ponds are well
known, including their effects on ecosystem functions, such as methane emissions (see
Chapters 4 and 17: “Wedand Hydrology” and “Wetlands and Climate Change”).

Birds

Waterfowl are plentiful in almost all wetlands, probably because of the food rich-
ness and the diversity of habitats for nesting and resting. Migratory waterfowl nest
in northern freshwater marshes, winter in southern marshes, and rest in other marshes
during their migrations. In a typical freshwater marsh, different species distribute
themselves along an elevation gradient according to how well they are adapted to
water (Fig. 10.7). In northern marshes, the loon {Gavia immer) usually uses the
deeper water of marsh ponds, which may hold fish populations. Grebes {Podilymbus
sp. and Podiceps sp.) prefer marshy areas, especially during the nesting season. Some
ducks (dabblers), such as mallards {Anas platyrhynchos), nest in upland sites, feeding
along the marsh-water interface and in shallow marsh ponds. Others (diving ducks),
such as the ruddy duck {Oxyura jamaicensis), nest over water and fish by diving. For
example, the black duck {Anas rubripes), one of the most popular ducks for naturalists
and hunters alike, uses the emergent marsh as its preferred habitat. The northern shov-
eler {A. clypeata), the “whale of the waterfowl,” uses its large bill and laternal lamellae
to hlter plankton. Geese {Branta canadensis and Chen sp.) and swans {Cy^nussp.),
the “cattle of the waterfowl,” along with canvasback ducks {Aythya valisineria) and
the wigeon {Anas americana), are major marsh herbivores. Wading birds, such as
the great blue heron {Ardea herodias) and the great egret {Casmerodius albus)^ usu-
ally nest colonially in wetlands and fish along the shallow ponds and streams. The
least bittern {Ixobrychus exilis) builds nests a meter or less above the water in Typha
or Scirpus/Schoenoplectus stands. Rails live in the whole range of wetlands; many of
them are solitary birds that are seldom seen. Marsh wrens {Cistohorus plaustris), Vir-
ginia rails {Rallus limicola) , soras {Porzana Carolina)^ and swamp sparrows {Melospiza
^eorpfiana) live amid the dense vegetation of freshwater marshes, often heard but not

358 Chapter 10 Freshwater Marshes

Figure 10.7 Typical distribution of birds across a freshwater marsh from open water edge
across shallow water to upland grasses. Placement of muskrat and mink is also illustrated.
(After Weller and Spatcher, 1965)

seen. Songbirds are also abundant in and around marshes. They often nest or perch in
adjacent uplands and fly into the marsh to feed. Swallows {Riparia riparia and Steljji-
dopteryx serripmnis) and swifts (e.g., chimney swifts Chaetura pdapfica) are common
around freshwater marshes, flying above the marsh, with their mouths ever open to
capture emerging insects, often in swarms of dozens or even hundreds of birds.

One of the most conspicuous wetland birds in United States freshwater marshes
is the blackbird, represented by the yellow-headed blackbird {Xcmthocephalus
xanthocephalus) in parts of the midwestern United States and the red-winged
blackbird {Apflaius phoniceus) in the eastern United States. The red-winged blackbird
is a very social species and is quite territorial, especially during the nesting season.

Ecosystem Function

Primary Productivity

The above-ground net primary productivity of inland marshes has been reported in a
number of studies (Table 10.3). Estimates are generally quite high, ranging upward

Ecosystem Function 359

Table 10.3 Selected primary production estimates for inland freshwater marshes

Dominant

Species

Location

Net Primary
Productivity
(g m-2 yri)

Reference

Reeds and Grasses

G/yceria maxima

Lake, Czech

900-4,3003

Kvet and Husak (1978)

Phragmites

Republic
Lake, Czech

1,000-6,000®

Kvet and Husak (1978)

communis
R communis

Republic

Denmark

1,400®

Anderson (1976)

Panicum hemitomon

Fioating coastal

1,700'’

Sasser et al. (1982)

Schoenoplectus

marsh, Louisiana
Lake, Czech

1,600-5,500®

Kvet and Husak (1978)

lacutsths

Sparganium

Republic

Prairie pothoie, iowa

1,066"

van der Valk and Davis (1978)

eurycarpum
Typha glauca

Prairie pothoie, iowa

2,297"

van der Vaik and Davis (1978)

T. latifolia

Oregon

2,040-2,210®

McNaughton (1966)

Typha spp.

Lakeside, Wisconsin

3,450®

Kiopatek (1974)

Typha spp.

centrai Ohio created

627±75" planted

Mitsch et ai. (2012)

Sedges and Rushes

Carex atheroides

marshes (12 yrs;
starting 3 years
after pianting)

Prairie pothoie, iowa

wetland; 772±9l"
unplanted wetland

2,858"

van der Vaik and Davis (1978)

Carex lacustris

Sedge meadow.

1,078-1,741®

Bernard and Soisky (1977)

Juncus effusus

New York
South Caroiina

1,860®

Boyd (1971)

Scirpus fluviatilis

Prairie pothoie, iowa

943®

van der Vaik and Davis (1978)

Broad-Leaved Monocots

Acorus calamus Lake, Czech

500-1,100®

Kevt and Husak (1978)

Republic

®Above- and below-ground vegetation.
'’Above-ground vegetation.

from about 1,000 g irT^ . Some of the best estimates, which take into account
underground production as well as that above-ground, come from studies of fishponds
in the Czech Republic. (These are small artificial lakes and bordering marshes used
for fish culture.) These estimates, some indicating values of over 6,000 g are

high compared with most of the North American estimates and even higher than the
productivity of intensively cultivated farm crops.

The emergent monocotyledons Phra^mites and Typha, two of the dominant
plants in freshwater marshes, have high photosynthetic efficiency. For Typha, effi-
ciency is highest early in the growing season, gradually decreasing as the season

360 Chapter 10 Freshwater Marshes

Figure 10.8 Relationship between highest mean monthly temperature and above-ground
standing crop of various sedges in freshwater wetiands and uplands. Data points are for
wetlands except where noted otherwise. (After Gorham, 1974)

progresses. Phra^mites, in contrast, has a fairly constant efficiency rate throughout
most of the growing season. The efficiencies of conversion by these plants in optimum
environments of 4 to 7 percent of photosynthetically active radiation are comparable
to those calculated for intensively cultivated crops such as sugar beets, sugarcane,
and corn.

Productivity variation is undoubtedly related to a number of factors, including
summer air temperatures (Fig. 10.8). Innate genetic differences among species
account for part of the variability. For example, in one study that used the same
techniques of measurement (Kvet and Husak, 1978), Typha anpustifolm production
was determined to be double that of T. latifoUa.

The dynamics of underground growth are much less studied than those of
aboveground growth. Annuals generally use small amounts of photosynthate to
support root growth, whereas species with perennial roots and rhizomes often have
root : shoot ratios well in excess of 1. This relationship also appears to hold true
tor inland freshwater marshes. Perennial species in freshwater marshes generally
have more below-ground than above-ground biomass (Fig. 10.9). Even when
biomass root : shoot ratios are usually greater than 1, ratios of root production to
shoot biomass are generally less than 1. Since above-ground production is often
approximated by above-ground biomass, this latter ratio is an index of the allocation
of resources by the plant, and it indicates that less than one-half of the photosynthate
is translocated to the roots. The coexistence of large root biomass and relatively small

Ecosystem Function 361

aboveground biomass
belowground biomass
I I lemnid biomass

-150

-50

Riragmitescommunis

Typha anguall folia

100 i

Figure 10.9 Distribution of above-ground and below-ground biomass of emergent vegeta-
tion and lemnids across a reed bed {Phragmites) transect, showing relation to elevation and
flooding. (After Kvet and Husak, 1978)

root production suggests that the root system is generally longer lived (i.e., it renews
itself more slowly) than the shoot.

Decomposition and Herbivory

With some notable exceptions, such as muskrat and geese grazing, herbivory is consid-
ered fairly minor in mland marshes where most of the organic production decomposes
before entering the detrital food chain. The decomposition process is much the same
for all wetlands. Variations stem from the quality and resistance of the decomposing
plant material, the temperature, the availability of inorganic nutrients to microbial
decomposers, and the flooding regime of the marsh.

Consumers play a significant role in detrital cycles. Most litter decomposition
studies in freshwater marshes were done with senesced plant material during the win-
ter, and generally low {k = 0.002-0.007 day^'^) rates were measured. However, in a
comparison of the decay of fresh biomass and senesced wetland plant leaves. Nelson
et al. (1990a,b) found that samples of freshly harvested wetland plant material (Typha
jjlauca) decomposed more than twice as fast (k = 0.024 day“^ ) as did naturally senesced
material (^ = 0.011 day“^ ). This comparison illustrates the more rapid decomposition
that results when animals such as muskrats harvest live plant material.

Muskrats also may play a positive role in the energy flow of a marsh as they harvest
aquatic plants and standing detritus for their muskrat mounds. Wainscott et al. (1990)

362 Chapter 10 Freshwater Marshes

found in culturing experiments that litter from muskrat mounds supports substantially
higher densities of microbes than litter from the marsh floor does. They suggested that
these muskrat mounds may act like “compost piles,” as they accelerate the decompo-
sition and microbial growth that have become familiar to organic gardeners.

Muskrat eat-outs and the resulting open water contribute to structural and bio-
geochemical heterogeneity within marshes. In an Iowa prairie pothole marsh, Rose
and Crumpton (2006) found that as vegetated zones transitioned into open water,
there was a predictable decrease in dissolved oxygen and increase in methane con-
centrations. They suggested that these conditions were regulated by the presence of
emergent vegetation and its influence on aerobic and anaerobic metabolism.

Food Webs

Even though food chains begin in the detrital material of freshwater marshes, they
develop into detailed webs that are still poorly understood. Benthic communities
that feed on detritus form the basis of food for fish and waterfowl in the marshes.
DeRoia and Bookhout (1989) found that chironomids made up 89 percent of the
diet of blue-winged teal {Anas discors) and 99 percent of the diet of green-winged
teal (A. crecca) in a Great Lakes marsh. The direct grazing of freshwater marsh veg-
etation has occasionally been reported in the literature. Crayfish are often important
consumers of macrophytes, particularly of submersed aquatic plants, in freshwater
marshes. The red swamp crayfish {Procambarus clarkii) was shown to have effec-
tively grazed on Potamo0eton pectinatm in a freshwater marsh in California, where
the plant decreased from 70 to 0 percent cover of the marsh while the crayfish popu-
lation almost doubled (Feminella and Resh, 1989). The direct consumption of marsh
plants by geese, muskrats, and other herbivores is common in some parts of the world.
Eat-outs causing large expanses of open water are the result of the inability of plants
to survive after being clipped below the water surface by animals (Middleton, 1999).

In one study that disputed the low-herbivory assumption of inland marshes,
deSzalay and Resh (1997) found that herbivores represented about 27 percent of
the benthic community in brackish inland marshes. These herbivores fed primarily
on filamentous algae and diatoms. Yet herbivory on marsh macrophytes generally
remains low except for large -animal grazing from time to time.

Nutrient Budgets

Vegetation traps nutrients in biomass, but the storage of these nutrients is seasonally
partitioned in above-ground and below- ground stocks. For example, nutrient stocks
in the roots and rhizomes of macrophytes are mobilized into the shoots early in the
growing season and increase to as much as 4 g P rcT^ during the summer. In the fall,
some nutrients in the shoots are translocated into the below-ground organs before
the shoots die, but most nutrients are lost by leaching and in the litter. Nitrogen and
carbon budgets for two created flowthrough marshes in Ohio dominated by Typha
spp. were presented in Chapter 6: “Wetland Biogeochemistry” (Fig. 6.19). In these

Ecosystem Function 363

Nitrogen Cycle Phosphorus Cycle

Figure 10.10 Fluxes ef nitrogen and phosphorus through a river bulrush (Scirpus fluviatilis)
stand in Wisconsin. Fiows are in g y~^, and storages are in g/m^ of nitrogen and phos-
phorus, respectively. (From Klopatek, 1978)

flowthrough wetlands, hydrologic fluxes dominated the wetlands (inflow, outflow, and
subsurface seepage), but the sequestration of nitrogen and carbon into the soil was one
of the largest fluxes. The nitrogen and phosphorus budgets for a freshwater Scirpus
marsh in Wisconsin (Fig. 10.10) show a peak biomass storage of 20.7 g N/m^ and
5.3 g P/m^. This plant storage is small compared to the nutrients that are stored
within the root zone of the peat and mineral soils (shown in Fig. 10.10 to be 1,700 g
N/m^ and 12 g P/m^ for total nitrogen and available phosphorus, respectively).

Studies such as these lead to five generalizations about nutrient cycling in fresh-
water marshes:

1 . The size of the plant stock of nutrients in freshwater marshes varies widely in
contrast to the much more abundant storage of nutrients in marsh soils.

More nitrogen and phosphorus are retained in above-ground plant parts in
mineral substrate wetlands (freshwater marshes) than in peadands due to the
higher productivity and higher concentrations of nutrients. The
above-ground stock of nitrogen ranges from as low as 3 to as high as 30 g-N
m~^ in freshwater marshes.

2. The biologically inactivated stock of nutrients in plants is only a temporary
storage that is released to flooding waters and sediments when the plant
shoots die in autumn. Where this occurs, the marsh may retain nutrients
during the summer and release them in the winter.

364 Chapter 10 Freshwater Marshes

3. Nutrients retained in biomass are often a small portion of nutrients that flow
into the marshes, and that percentage decreases with increased nutrient
input. Thus, as more nutrients become available to a freshwater marsh, the
marsh becomes more “leaky.” Nutrients are lost from the system, and
nutrient turnover in the vegetation increases. Even if the uptake rate of
nutrients is high in wetlands, some of those nutrients are returned via detrital
decomposition to the nutrient pool in the sediments and overlying waters. If
wetlands are used for nutrient removal (see Chapter 19 “Wetlands and Water
Quality”), then it is common for only 10 to 20 percent of the nutrient inflow
to transfer temporarily into plant biomass.

4. Marsh vegetation often acts as a nutrient pump, taking up nutrients from the
soil, translocating them to the shoots, and releasing them on the marsh
surface during senescence. The effect of this pumping mechanism may be to
mobilize nutrients that have been sequestered in the soil. In some cases, the
uptake of nutrients by macrophytes from the sediments is considerably higher
than the inflow. Most of this uptake is translocated back to the roots or lost
through leaching and shoot senescence, so biomass storage of nutrients is
generally low compared to annual inflow.

5. In general, precipitation and dryfall account for less than 10 percent of plant
nutrient demands in freshwater marshes. Similarly, groundwater flows are
usually small sources of phosphorus, but, in agricultural settings with artificial
drainage, nitrate-nitrogen inflow can be high. Surface inflow is usually a major
source of phosphorus because of its ability to sorb onto sediments, particularly
clay. Considering all of these variables, it is not surprising that each marsh
seems to have its own unique nutrient budget. In low-nutrient wetlands like
the Florida Everglades, the marsh system is accustomed to relying primarily
on nutrient inflow from precipitation and dry fallout from hres.

Nutrient Limitations

Koerselman and Meuleman (1996), in a study of several wetlands in Europe, found
that the nitrogen iphosphorus (N:P) ratios in wetland plant tissues were correlated
with the N:P supply ratio and that any N:P ratio less than 14:1 suggests nitrogen
limitation. This is twice the often-used Redfield ratio (N:P = 7.2 by weight) that is used
in planktonic systems to indicate relative nutrient limitation. As part of an extensive
literature review on temperate North American wetlands, Bedford et al. (1999) found
that only marshes were consistently N-Umited as indicated by leaf- tissue and soil N:P
ratios <14 (although swamp tended to have soils with N:P ratios <14 as well). Based
on their leaf tissue N:P ratio, other wetland types (swamps, bogs, fens) tended to be
co-limited by N and P or just P-limited based on the thresholds derived by Koerselman
and Meuleman (1996).

Mcjannet et al. (1995) investigated the nitrogen and phosphorus content of 41
freshwater marsh plants after they were grown in excess fertilizer for one growing
season. There was a wide range of nitrogen (0.25-2.1 percent N) and phosphorus

Ecosystem Function 365

(0.13-1.1 percent P) that was not related to where the plants came from. However,
plants that were from ruderal life histories (i.e., annuals or functional annuals) did have
signifrcandy lower nitrogen and phosphorus tissue concentrations than did perennials.

For a Manitoba Scirpus acutus marsh, Neill (1990) found that neither nitrogen
nor phosphorus increased net productivity when applied alone but that above-ground
biomass nearly doubled when nitrogen and phosphorus were applied together. Similar
studies of a nearby marsh showed nitrogen limitation, indicating that differences in
limiting factors are possible even in the same region (Neill, 1990). Under conditions
in which water levels are more stable, such as Louisiana’s Gulf Coast, the addition of
nitrogen fertilizer at a rate of 10 g NH4'*‘-N/m^ caused approximately a 100 percent
increase in the growth of Sagittaria Icmcifolm (Delaune and Lindau, 1990).

Experiments by Svengsouk and Mitsch (2001) support the multiple -nutrient limi-
tation of some freshwater marsh plants. Their study investigated the relative limitations
of nitrogen and phosphorus in mesocosms planted with both bulrush {Schoenoplec-
tus tabernaemontani) and cattail (Typha sp.) together. Results suggested that, when
both nitrogen and phosphorus are available, Typha competed well with Schoenoplec-
tus. When only one of the nutrients was in abundance, Schoenoplectus did much better
than Typha.

In contrast, enrichment studies by Craft et al. ( 1995 ) on the low-nutrient sawgrass
( Cladium jamaicmse) and other macrophyte communities illustrated that the most
important limiting factor in the Florida Everglades is phosphorus. Nitrogen additions
had no effect on biomass production, nutrient uptake rates, or nitrogen enrichment
of peat. Phosphorus enrichment from agricultural sources has attributed to substantial
ecological change in the Everglades, most notably the transition of large areas from
sawgrass to cattail ( Typha spp.). Restoring the Everglades requires reducing P-enriched
agricultural stormwater entering the region. (See Chapter 19: “Wetlands and Water
Quality,” Case Study 2).

Greenhouse Gas Emissions

The anaerobic conditions in marshes and other wetlands give them the potential to
emit considerable amounts of nitrous oxide (N2O) and methane (CH4) both of which
are considered important greenhouse gases. Temperature and diffusion rates through
water influence the net emission of these gases, and therefore shallow wetlands often
have greater emissions than open water bodies. In a boreal lake in Finland, Huttunen
et al. (2003) estimated that the littoral zone, which consisted of 26 percent of the total
lake surface area, was responsible for most of the N2O emissions from the lake. Brix
et al. (2001) examined whether a Phrapimites mz.Kh. in Europe could be considered a
net source or sink for greenhouse gases given that marshes assimilate carbon dioxide
(CO2) and store carbon while emitting CH4. They found that when these marshes are
evaluated over a shorter time period (<60 years), these wedands could be considered
a net source of greenhouse gases based on their emission of CH4 and CO2 relative
to carbon ftxation. However, CH4 does not persist in the atmosphere indefinitely.
If these marshes are evaluated over a longer time period (>100 years), the balance

366 Chapter 10 Freshwater Marshes

shifts and the marshes are a net a sink for greenhouse gases. Mitsch et al. (2013)
found similar results in a comparison of 21 wetlands from around the world, many of
them freshwater marshes. Several recent methane emissions studies from created and
natural wetlands in Ohio and Costa Rica by Altor and Mitsch (2006, 2008), Nahlik
and Mitsch (2010, 2011), Sha et ah, (2011), Mitsch et al. (2013), and Waletzko and
Mitsch (2014) suggest that methane emissions from created wetlands are lower in
their first two decades than are comparable natural reference wetlands.

Recommended Readings

Coburn, E. A. 2004. Vernal Pools . Blacksburg, VA: McDonald & Woodward,
van der Valk, A. G. 2012. The Biolojjy of Freshwater Wetlands, 2nd ed. Oxford, UK:
Oxford University Press

References

Altor, A. E., and W. J. Mitsch. 2006. Methane flux from created riparian marshes:
Relationship to intermittent versus continuous inundation and emergent macro-
phytes. Ecolo£iical Engineering 2B\ 224-234.

Altor, A. E., and W. J. Mitsch. 2008. Pulsing hydrology, methane emissions, and car-
bon dioxide fluxes in created marshes: A 2-year ecosystem study. Wetlands 28:
423-438.

Anderson, F. O. 1976. Primary productivity in a shallow water lake with special refer-
ence to a reed swamp. Oikos 27: 243-250.

Balogh, G. R., and T. A. Bookhout. 1989. Purple loosestrife (Lythrum salicaria) in
Ohio’s Lake Erie marshes. Ohio Journal of Science 89: 62-64.

Bedford, B. L., M. R. Walbridge, and A. Aldous. 1999. Patterns of nutrient availability
and plant diversity of temperate North American wetlands. Ecology 8:125 1-1269 .

Bernard, J. M., and B. A. Solsky. 1977. Nutrient cycling in a Carex /acMttw wetland.
Canadian Journal of Botany 55: 630-638.

Boyd, G. E. 1971. The dynamics of dry matter and chemical substances in a Juncus
population. American Midland Naturalist 86: 28M:5.

Bridgham, S. D., K. Updegraff, and J. Pastor. 2001. A comparison of nutrient avail-
ability indices along an ombrotrophic-minerotrophic gradient in Minnesota wet-
lands. Soil Science Society of America Journal 65: 259-269.

Brix, H., B. K. Sorrell, and B. Lorenzen. 2001. Are Phragmites-dominated wedands
a net source or net sink of greenhouse gases.^ Aquatic Botany 69: 313-324.

Graft, G. B., J. Vymazal, and C. J. Richardson. 1995. Response of Everglades plant
communities to nitrogen and phosphorus additions. Wetlands 15: 258-271.

Dahl, T. E. 2006. Status and trends of wetlands in the conterminous United States 1 998
to 2004. U.S. Department of the Interior, Fish and Wildlife Service, Washington,
DC, 112 pp.

Delaune, R. D., and C. W. Lindau. 1990. Fate of added 15N labeled nitrogen in
a Sa^ittaria lancifolia L. Gulf Coast marsh. Journal of Freshwater Ecolojyy 5:
429-431.

References 367

Denny, P. 1985. Wetland plants and associated plant life-forms. In P. Denny, ed.. The
Ecolojjy and Manajjement of Wetland Ve£ietation. Junk, Dordrecht, the Nether-
lands, pp. 1-18.

Denny, P. 1993. Wedands of Africa: Introduction. In D. F. Whigham, D. Dykyjova,
and S. Hejny, eds. Wetlands of the World, I: Inventory, Ecology, and Mana£fement .
Kluwer Academic Publishers, Dordrecht, the Netherlands, pp. 1-31.

Derksen, A. J. 1989. Autumn movements of underyearling northern pike, Esox lucius,
from a large Manitoba marsh. Canadian Eield Naturalist 103:429-431.

DeRoia, D. M., and T. A. Bookhout. 1989. Spring feeding ecology of teal on the Lake
Erie marshes (abstract). Ohio Journal of Science 89(2): 3.

deSzalay, F. A., and V. H. Resh. 1997. Responses of wedand invertebrates and plants
important in waterfowl diets to burning and mowing of emergent vegetadon.
Wetlands 17: 149-156.

Farrar, J., and R. Gersib. 1991. Nebraska salt marshes: Last of the least. Nebraska
Game and Park Gommission, Lincoln. 23 pp.

Feminella, J. W., and V. H. Resh. 1989. Submersed macrophytes and grazing crayfish:
An experimental study of herbivory in a California freshwater marsh. Holarctic
Ecology 12: 1-8.

Galatowitsch, S. M., N. O. Anderson, and P. D. Ascher. 1999. Invasiveness in wetland
plants in temperate North America. Wetlands 19: 733-755.

Gorham, E. 1974. The relationship between standing crop in sedge meadows and
summer temperature. Journal ofEcolojyy 62: 487M91.

Hall, J. V., W. E. Frayer, and B. O. Wilen. 1994. Status of Alaska wetlands. U.S. Fish
& Wildlife Service, Alaska Region, Anchorage. 32 pp.

Huttunen, J. T., H. Nykanen, J. Turunen, and P. J. Martikainen. 2003. Methane emis-
sions from natural peatlands in the northern boreal zone in Finland, Fennoscan-
dia. Atmospheric Environment 37: 147-151.

Johnson, S. A. 1998. Effects of hydrology and plant introduction techniques on
first-year macrophyte growth in an newly created wedand. Master’s thesis, Ohio
State University, Columbus.

Johnston, C. A., S. D. Bridgham, and J. P. Schubauer-Berigan. 2001. Nutrient dynam-
ics in relation to geomorphology of riverine wetlands. Soil Science of America
Journal 65: 557-577.

Johnston, C. A., B. Bedford, M. Bourdaghs, T. Brown, C. Frieswyk, M. Tulbere, L.
Vaccaro, and J. B. Zedler. 2007. Plant species indicators of physical environment
in Great Lakes coastal wedands. Journal of Great Lakes Research 33: 106-124.

Keddy, P. A. 2010. Wetland Ecology: Principles and Conservation, 2nd ed. Cambridge
University Press, Cambridge, UK. 497 pp.

Klopatek, J. M. 1974. Producdon of emergent macrophytes and their role in min-
eral cycling within a freshwater marsh. Master’s thesis. University of Wisconsin,
Milwaukee.

Klopatek, J. M. 1978. Nutrient dynamics of freshwater riverine marshes and the role of
emergent macrophytes. In R. E. Good, D. F. Whigham, and R. L. Simpson, eds..

368 Chapter 10 Freshwater Marshes

Freshwater Wetlands: Ecological Processes and Management Potential. Academic
Press, New York, pp. 195-216.

Koerselman, W., and A. F. M. Meuleman. 1996. The vegetation N:P ratio: A new
tool to detect the nature of nutrient limitation. Journal of Applied Ecology 33:
1441-1450.

Kulesza, A. E., and J. R. Holomuzki. 2006. Amphipod performance responses to
decaying leaf litter of Phra^mites australis and Typha an^ustifolia from a Lalce
Erie coastal marsh. Wetlands 26: 1079-1088.

Kvet, J., and S. Husak. 1978. Primary data on biomass and production estimates in
typical stands of fishpond littoral plant communities. In D. Dykyjova and J. Kvet,
eds.. Pond Littoral Ecosystems. Springer-Verlag, Berlin, pp. 211-216.

LaBaugh, J.W. 1989. Chemical characteristics of water in northern prairie wetlands.
In A. van der Valk, ed., Northern Prairie Wetlands. Iowa State University Press,
Ames, lA.

LaBaugh, J. W., T. C. Winter, G. A. Swanson, D. O. Rosenberry, R. D. Nelson, and N.
H. Euliss. 1996. Changes in atmosphereic circulation patterns affect midcontinent
wetlands sensitive to climate. Limnology & Oceanography 41: 864-870.

Matthews, E., and 1. Fung. 1987. Methane emissions from natural wetlands: Global
distribution, area, and environmental characteristics of sources. Global Bio^eo-
chemical Cycles 1: 61-86.

Mcjannet, C. L., P. A. Keddy, and F. R. Pick. 1995. Nitrogen and phosphorus tissue
concentrations in 41 wetland plants: A comparison across habitats and functional
groups. Functional Ecology 9: 231-238.

McLaughlin, D. B., and H. J. Harris. 1990. Aquatic insect emergence in two Great
Lakes marshes. Wetlands Ecology and Management 1: 111-121.

McNaughton, S. J. 1966. Ecotype function in the Typha community-type. Ecological
Monographs 36: 297-325.

Middleton, B. A. 1999. Wetland Restoration: Flood Pulsing and Disturbance Dynam-
ics. John Wiley & Sons, New York, 388 pp.

Mitsch, W. J., L. Zhang, K. C. Stefanik, A. M. Nahlik, C. J. Anderson, B. Bernal,
M. Hernandez, and K. Song. 2012. Creating wetlands: Primary succession, water
quality changes, and self-design over 15 years. BioScience 62: 237-250.

Mitsch, W. J., B. Bernal, A. M. Nahlik, U. Mander, L. Zhang, C. J. Anderson, S. E.
Jorgensen, and H. Brix. 2013. Wetlands, carbon, and climate change. Landscape
Ecology 28: 583-597.

Moore, D. R. J., P. A. Keddy, C. L. Gaudet, and 1. C. Wisheu. 1989. Conservation of
wetlands: do infertile wetlands deserve a higher priority.^ Biological Conservation
47: 203-217.

Nahlik, A. M., and W. J. Mitsch. 2010. Methane emissions from created riverine wet-
lands. Wetlands 30: 783-793. Erratum in Wetlands {2QW) 31: 449-450.

Nahlik, A. M., and W. J. Mitsch. 2011. Methane emissions from tropical freshwater
wetlands located in different climatic zones of Costa Rica. Global Chanpfe Biology
17: 1321-1334.

References 369

Neill, C. 1990. Nutrient limitation of hardstem bulrush {Scirpus acutus Muhl.) in a
Manitoba interlake region marsh. Wetlands 10: 69-7 S.

Nelson, J. W., J. A. Kadlec, and H. R. Murkin. 1990a. Seasonal comparisons of weight
loss of two types of Typha ^lauca Godr. leaf litter. Aquatic Biology 37: 299-314.

Nelson, J. W., J. A. Kadlec, and H. R. Murkin. 1990b. Response by macroinvertebrates
to cattail litter quality and timing of litter submergence in a northern prairie marsh.
Wetlands 10: 47-60.

Philipp, K. R., and R. T. Field. 2005. Bhragmites australis in Delaware Bay

salt marshes. Ecological Enpfineerinpi 275-291.

Porej, D. 2004. Faunal aspects of wetland creation and restoration. Ph.D. dissertation.
Evolution, Ecology, and Organismal Biology Department, Ohio State University,
Columbus.

Richter, K. O., and A. L. Azous. 1995. Amphibian occurrence and wetland charac-
teristics in the Puget Sound Basin. Wetlands 15: 305-312.

Rose, C., and W. G. Crumpton. 2006. Spatial patterns in dissolved oxygen and
methane concentrations in prairie pothole wetlands in Iowa, USA. Wetlands 26:
1020-1025.

Saltonstall, K. 2002. Cryptic invasion by a non-native genotype of the common reed,
Phrqgmites australis^ into North America. Proceedings of the National Academy of
Sciences 99: 2445-2449.

Saltonstall, K., P. M. Peterson, and R. J. Soreng. 2004. Recognition of Phrqgmites aus-
tralis suhsp . americanus (Poaceae: Arundinoideae ) in North America: Evidence
from morphological and genetic analyses. Brit. Org/SIDA 21: 683-692.

Sasser, C. E., G. W. Peterson, D. A. Fuller, R. K. Abernethy, and J. G. Gosselink.
1982. Environmental Monitoring Program, Louisiana Offshore Oil Port Pipeline,
1981. Annual Report, Coastal Ecology Laboratory, Center for Wedand Resources,
Louisiana State University, Baton Rouge, 299 pp.

Sha, C., W. J. Mitsch, U. Mander, J. Lu, J. Batson, L. Zhang and W. He. 2011.
Methane emissions from freshwater riverine wetlands. Ecological Engineering 37:
16-24.

Stephenson, T. D. 1990. Fish reproductive utilization of coastal marshes of Lake
Ontario near Toronto. Journal of Great Lakes Research 16: 71-81.

Steward, K. K. 1990. Aquatic weed problems and management in the eastern United
States. In A. H. Pietersen, and K. J. Murphy, eds. Aquatic Weeds: The Ecology and
Management of Nuisance Aquatic Vegetation. Oxford University Press, New York,
pp. 391-405.

Stuckey, R. L. 1980. Distributional history of Lythrum salicaria (purple loosestrife)
in North America. Bartonia 47: 3-20.

Svengsouk, L. M., and W. J. Mitsch. 2001. Dynamics of mixtures of Typha latifolia
and Schoenoplectus tabernaemontani in nutrient-enrichment wetland experiments.
American Midland Naturalist 145: 309-324.

van der Valk A. G., and C. B. Davis. 1978. Primary production of prairie glacial
marshes. In: Ereshwater Wetlands: Ecological Processes and Management Potential,

370 Chapter 10 Freshwater Marshes

R. E. Good, D. F. Whigham, and RL. Simpson, eds.. Academic Press, New York,
pp. 21-37.

Wainscott, V. J., C. Bardey, and P. P Kangas. 1990. Effect of muskrat mounds on
microbial density on plant litter. American Midland Naturalist 123: 399-401.

Waletzko, E., and W. J. Mitsch. 2014. Methane emissions from wetlands: An in situ
side-by-side comparison of two static accumulation chamber designs. Ecological
Engineering 71\ 95-102.

Weller, M. W. 1994. Freshwater Marshes^ 3rd ed. University of Minnesota Press, Min-
neapolis. 192 pp.

Weller, M. W., and C. S. Spatcher. 1965. Role of habitat in the distribution and abun-
dance of marsh birds. Iowa State University Agriculture and Home Economics
Experiment Station Special Report 43, Ames, lA, 31 pp.

Werner, K. J., and J. B. Zedler. 2002. How sedge meadow soils, microtopography,
and vegetation respond to sedimentation. Wetlands 22: 451M66.

Zoltai, S. C. 1988. Wetland environments and classification. In National Wetlands
Working Group, ed.. Wetlands of Canada. Ecological Land Glassification Series
24, Environment Ganada, Ottawa, Ontario, and Polyscience Publications, Mon-
treal, Quebec, pp. 1-26.

Freshwater swamp in Florida

Chapter H

Freshwater Swamps and Riparian
Ecosystems

Freshwater forested wetlands in North America ran£;e from deepwater swamps
dominated by bald cypress-tupelo (Taxodium distichum-Nyssa aquatica^,
pond cypress-black ^um (Taxodium distichum var. imbricarium-Nyssa
sylvatica var. biflora^^ and Atlantic white cedar (Chamaecyparis thyoides^
swamps found alonp; the eastern seaboard of the United States, to less wet red
maple (Acer rubrum^ swamps found throup/hout New England and the
Mid-Atlantic states. Riparian ecosystems have soils and soil moisture influenced
by the adjacent stream or river and are unique because of their linear form
alonp! rivers and streams and because they process lar^e fluxes of enerpyy and
materials from upstream systems. Riparian ecosystems include bottomland
hardwood forests found alon^ rivers in mesic climates everywhere. Trees in
forested wetlands have developed several unique adaptations to the wetland
environment, includinpi knees, wide buttresses, adventitious roots, fluted trunks,
andpfas transport to the rhizosphere. Forested swamp primary productivity is
closely tied to hydrolopfic conditions with lower productivity whenever conditions
are either too wet or too dry. The function of riparian ecosystems is much better
explained by a generalized theory called the flood pulse concept than by a
previous theory of streams referred to as the river continuum concept.

In the nomenclature used in this book, swamps are forested wetlands. We discussed
saltwater swamps in Chapter 9: “Mangrove Swamps.” There are an estimated 1.1
million km^ of freshwater swamps in the world, representing about 20 percent of the
inland wedands of the world (Table 10.1).

Very few trees flourish in standing water. Exceptions are found in the southeastern
United States, where cypress {Taxodium sp.) and tupelo/gum {Nyssa sp.) swamps

373

374 Chapter 11 Freshwater Swamps and Riparian Ecosystems

are found in deepwater forested wedands and are characterized by bald cypress-water
tupelo communities with permanent or near-permanent standing water. These
so-called deepwater swamps were defined by Penfound (1952) as having “fresh water,
woody communities with water throughout most or all of the growing season” and
include isolated cypress domes and alluvial cypress swamps along rivers. Along the
middle-eastern seaboard of the United States and along the Florida Panhandle, the
cypress swamp partially gives way to another forested wetland, the Atlantic white
cedar [Chamaecyparis thyoides) swamp. Farther northeast through New England and
well into the Midwest, other types of freshwater forested wetlands occur, although
they are not as wet as the cypress-tupelo swamps, nor are the tree species coniferous
as are cypress. These broad-leaved deciduous forested wetlands include forest found
along river floodplains (riparian forests or bottomland hardwood forests) and a
multitude of forested wetlands that are found in isolated upland depressions.

Extensive tracts of riparian wetlands, which occur along rivers and streams, are
occasionally flooded by those bodies of water but are otherwise dry for varying por-
tions of the growing season. Riparian forests and freshwater swamps combined con-
stitute the most extensive class of wetlands in the United States, covering an estimated
280,000 km^. In the southeastern and midwestern United States, riparian ecosystems
are often referred to as bottomland hardwood forests. They contain diverse vegetation
that varies along gradients of flooding frequency. Riparian wetlands also occur in arid
and semiarid regions of the United States, where they are often a conspicuous feature
of the landscape in contrast to the surrounding arid grasslands and desert. Riparian
ecosystems are generally considered to be more productive than the adjacent uplands
because of the periodic inflow of nutrients, especially when flooding is seasonal rather
than continuous.

Geographic Extent

Cypress-Tupelo Swamps

Bald cypress (Taxodium distichum [L.] Rich.) swamps are found as far north as south-
ern Illinois and western Kentucky in the Mississippi River embayment and southern
New Jersey along the Atlantic Coastal Plain in the United States (Fig. 1 1. 1 a). Pond
cypress [Taxodium distichum var. imbricarium [Nutt.] Groom), described variously
as either a different species or a subspecies of bald cypress (Denny and Arnold, 2007),
has a more limited range than bald cypress and is found primarily in Florida and
southern Georgia; it is not present along the Mississippi River floodplain except in
southeastern Louisiana. There is a third cypress, Montezuma cypress ( T. distichum var.
mexicanum Gordon), that is found in Mexico and southern Texas. Another species
indicative of the deepwater swamp is the water tupelo [Nyssa aquatica L.), which
has a range similar to that of bald cypress along the Atlantic Coastal Plain and the
Mississippi River, although it is generally absent from Florida except for the west-
ern peninsula. Water tupelo occurs in pure stands or is mixed with bald cypress in
floodplain swamps.

Geographic Extent 375

Figure 11.1 Distribution of dominant forested wetland trees in the seutheastern United
States: (a) bald cypress {Taxodium distichum) and pond cypress {Taxodium distichum var.
imbricarium) (with dotted line indicating nerthern extent of pond cypress) and (b) white
cedar (Chamaecyparis thyoldes). (After Little, 1971; Laderman, 1989)

White Cedar Swamp

White cedar swamps, dominated by Atlantic white cedar ( Chamaecyparis thyoides [L. ]
BSP), were once abundant along the Atlantic and Gulf of Mexico coastlines of the
United States as far north as southeastern Maine (Fig. 11.1b). These wetlands are
not nearly as plentiful as are cypress-tupelo swamps. White cedar occurs in about
2,150 km^ of forestland, but the species accounts for a majority of the trees in only
about 442 km^ (Sheffield et ah, 1998). Only 53 km^ of Atlantic white cedar swamps

376 Chapter 11 Freshwater Swamps and Riparian Ecosystems

remain in the glaciated northeastern United States, with red maple {Acer rubrum L.)
swamps now more prevalent there. The three states that have the most area of tim-
berland with Adantic white cedar are North Carolina, Florida, and New Jersey. The
regions with the highest concentrations of Atlantic white cedar are the Pinelands of
southeastern New Jersey; the Dismal Swamp of Virginia and North Carolina; and the
floodplains of the Escambia, Apalachicola, and Blackwater rivers in Florida (Sheffield
etal., 1998).

Red Maple Swamps

One of the most common of the broad-leaved deciduous forested wetlands in the
northeastern United States is the red maple {Acer rubrum) swamp. Toward the west
into Pennsylvania and Ohio, red maple swamps are replaced by swamps dominated
by trees such as ash {Fraxinus spp.), American elm ( Ulmus americana), swamp white
oak {Quercus bicolor)^ and a number of other species, but in the northeastern United
States, the red maple swamp is the most common swamp. Using an approximation
that all broad -leaved deciduous forested wetlands in several of the coastal states of the
northeastern United States are red maple swamps (this approximation would not apply
west or south of New York), Golet et al. (1993) estimated that there were 3,530 km^
of red maple swamps in these six states. Red maple forests also occur in the Upper
Peninsula of Michigan and northeastern Wisconsin. The range of the species Acer
rubrum extends westward to the Mississippi River and northward through much of
Ontario and parts of Manitoba and Newfoundland, but the tree can grow in both
wetlands and dry, sandy or rocky uplands. Thus, the presence of red maple does not
always indicate wetlands, as would the presence of cypress, tupelo, or white cedar.

Riparian Ecosystems

In general terms, riparian ecosystems are found wherever streams or rivers at least
occasionally cause flooding beyond their channel confines or where new sites for vege-
tation establishment and growth are created by channel meandering (e.g., point bars).
In arid regions, riparian vegetation may be found along or in ephemeral streams as
well as on the floodplains of perennial streams. In most nonarid regions, floodplains
and hence riparian zones tend to appear first along a stream “where the flow in the
channel changes from ephemeral to perennial — that is, where groundwater enters the
channel in sufficient quantity to sustain flow through nonstorm periods” (Leopold
et ah, 1964).

Riparian ecosystems can be broad alluvial valleys several tens of kilometers wide
or narrow strips of streambank vegetation in the arid regions. The “abundance of
water and rich alluvial soils” (Brinson et al., 1981) are the factors that make riparian
ecosystems different from upland ecosystems. Three major features separate riparian
ecosystems from other ecosystem types:

1 . Riparian ecosystems generally have a linear form as a consequence of their
proximity to rivers and streams.

Geomorphology and Hydrology 377

2. Energy and material from the surrounding landscape converge and pass
through riparian ecosystems in much greater amounts than those of any
other wetland ecosystem; that is, riparian systems are open systems.

3. Riparian ecosystems are functionally connected to upstream and downstream
ecosystems and are laterally connected to upslope (upland) and downslope
(aquatic) ecosystems.

Mesic Riparian Ecosystems

Mesic riparian ecosystems, commonly called bottomland hardwood forests or bottom-
land hardwoods in the United States, are one of the dominant types of riparian ecosys-
tems. Historically the term bottomland hardwood forest hiLS been used to describe the
vast forests that occur on river floodplains of the eastern and central United States,
especially in the Southeast. Bottomland hardwood forests are particularly notable wet-
lands because of the large areas that they cover in the southeastern United States and
because of the rapid rate at which they are being converted to other uses, such as agri-
culture and human settlements. This ecosystem is particularly prevalent in the lower
Mississippi River alluvial valley as far north as southern Illinois and western Kentucky
and along many streams that drain into the Atlantic Ocean on the south Atlantic
Coastal Plain. The Nature Conservancy (1992) estimated that before European set-
tlement, the Mississippi River alluvial plain supported about 21 million ha of riparian
forests; about 4.9 million ha remained as of 1991. The Atlantic Coastal Plain from
Maryland to Florida is another area of dense riparian forests lining the many rivers
that flow into the ocean.

Arid Riparian Ecosystems

Along high-order rivers, the contrast in elevation and vegetation between mesic ripar-
ian ecosystems and upland forests is often subtle and the gradients are gradual, whereas
for arid riparian ecosysems, the gradients are usually sharp and the visual distinctions
are usually clear. In the western United States and many other arid parts of the world,
these narrow riparian zones have been extensively modified by human activity. Conver-
sion to housing or agriculture is widespread. Damage from grazing animals is almost
ubiquitous. In an area where vegetation is generally limited by the lack of water, ripar-
ian vegetation and the availability of water inevitably draw and concentrate cattle.
Grazing along these primarily low-order streams results in increased erosion and chan-
nel downcutting while higher-order streams have been modified for water use.

Geomorphology and Hydrology

Cypress Swamps

Southern cypress-tupelo swamps occur under a variety of geologic and hydrologic
conditions, ranging from the extremely nutrient-poor dwarf cypress communities of
southern Florida to the rich floodplain swamps along many tributaries of the lower

378 Chapter 11 Freshwater Swamps and Riparian Ecosystems

Mississippi River basin. A useful classification of deepwater swamps in terms of their
geological and hydrological conditions includes five types (Fig. 11.2):

1. Cypress domes. Cypress domes (sometimes called cypress ponds or cypress
heads) are poorly drained to permanently wet depressions dominated by pond
cypress. They are generally small in size, usually 1 to 10 ha, and are numerous in the
upland pine flatwoods of Florida and southern Georgia. Cypress domes are found in
both sandy and clay soils and usually have several centimeters of organic matter that
has accumulated in the wetland depression. These wetlands are called domes because
of their appearance when viewed from the side: The larger trees are in the middle,
and smaller trees are toward the edges (Fig. 11.2a). Ewel and Wickenheiser (1988)
confirmed that trees grow slowest at the edges and fastest near the center of the
domes but found no significant differences in tree growth among small, medium,
and large cypress domes. This dome phenomenon, it has been suggested, is caused
by deeper peat deposits in the middle of the dome, fire that is more frequent around
the edges of the dome, or a gradual increase in the water level that causes the dome
to “grow” from the center outward (Vernon, 1947; Kurz and Wagner, 1953; Watts
et al., 2012). A definite reason for this profile has not been determined, nor do all
domes display the characteristic shape. An example of a water budget for a cypress
dome in north-central Florida is shown in Figure 11.3a.

2. Dwarf cypress swamps. Dwarf cypress swamps are major areas in southwestern
Florida, primarily in the Big Cypress Swamp and the Everglades, where pond cypress
is the dominant tree, but it grows stunted and scattered in a herbaceous understory
marsh (Fig. 11.2b). The trees generally do not grow more than 6 or 7 m high and
are more typically 3 m in height. The poor growing conditions are caused primar-
ily by the lack of suitable substrate overlying the bedrock limestone that is found in
outcrops throughout the region. The hydroperiod includes a relatively short period
of flooding as compared with other deepwater swamps, and fire often occurs. The
cypress, however, are rarely killed by fire because of the lack of fuel buildup and litter
accumulation.

3. Lake-edpie sivamps. Bald cypress swamps are also found as margins around
many lakes and isolated sloughs in southeastern United States, ranging from Florida
to southern Illinois (Fig. 11.2c). Tupelo and water-tolerant hardwoods such as
ash (Fraxinus spp.) often grow in association with the bald cypress. A seasonally
fluctuating water level is characteristic of these systems and is necessary for seedling
survival. The trees in these systems receive nutrients from the lake as well as from
upland runoff The lake-edge swamp can be a filter that receives overland flow from
the uplands and allows sediments to settle out and chemicals to adsorb onto the
sediments before the water discharges into the open lake. The importance of this
filtering function, however, has not been adequately investigated.

4. Slow-flowinp cypress strands. Cypress strands (Fig. 11. 2d) are found primar-
ily in southwest Florida, where the topography is slight, and rivers are replaced by
slow-flowing strands with little erosive power. The substrate is primarily sand, and

a Cypress dome

Map view of
Drainage

Dome shape

b. Dwarf cypress

c Lake-edge
swamp

d. Cypress strand

e. Alluvial river
swamp

Flood

Figure 11.2 General profile and flow pattern of major types of deepwater swamps, showing
(a) cypress dome, (b) dwarf cypress, (c) lake-edge swamp, (d) cypress strand, and (e) alluvial
river swamp. (After H. T. Odum, 1982)

379

Figure 11.3 Annual water budgets for (a) Florida cypress dome and (b) southern Illinois
cypress-tupelo alluvial cypress swamp, ((a) After Heimburg, 1984; (b) After Mitsch et al.,
1979)

380

Geomorphology and Hydrology 381

there is some mixture of limestone and remnants of shell beds. Peat deposits are shal-
low on higher ground and deeper in the depressions. The hydroperiod has a seasonal
wet-and-dry cycle. The deeper peat deposits usually retain moisture even in extremely
dry conditions. Much is known about cypress strands from many studies done in Faka-
hatchee Strand and Corkscrew Swamp (e.g., Carter et al., 1973; Duever et ah, 1984;
Villa and Mitsch, 2014, 2015).

5. Alluvial river swamps. The broad alluvial floodplains of rivers and creeks in
humid climates support a vast array of forested wetlands. In the southeastern and lower
Mississippi River basin, some of these are permanently flooded deepwater swamps as
part of a seasonally flooded forest (Fig. 11. 2e). Alluvial river swamps, dominated by
bald cypress or water tupelo or both in the southeastern United States, are confined
to permanently flooded depressions on floodplains such as abandoned river channels
( oxbows or billabon^s in Australia) or elongated swamps that usually parallel the river
{sloughs). Alluvial river swamps are continuously or almost continuously flooded. The
hydrologic inflows are dominated by runoff from the surrounding uplands and by
overflow from the flooding rivers. A water budget for an alluvial cypress-tupelo swamp
in southern Illinois is shown in Figure 11.3b, and a phosphorus budget for the same
swamp, showing the importance of the river input, is shown in Figure 6.18.

White Cedar Swamps

White cedar swamps occupy a narrow hydrologic niche generally between deepwater
cypress-tupelo swamps and moist-soil red maple swamps. The hydrologic regime of
cedar swamps can be classified as seasonally flooded, with flooding for an extended
period during the growing season. Golet and Lowry (1987) found that a group of
swamps in Rhode Island had a wide variability in annual water-level fluctuations, rang-
ing from 17 to 75 cm in amplitude and averaging 42 cm over a seven-year period.
The percentage of wetland flooded during the growing season ranged from 18 to
76 percent.

Red Maple Swamps

Red maple swamps and mineral-soil forested wedands occur, in general, in several
different hydrogeomorphic regimes, the most common being isolated basins in glacial
till or glaciofluvial deposits left behind by glaciations. The hydroperiod for two red
maple swamps in Rhode Island is shown in Figure 4.4, and the different hydrologic
settings for these types of wetlands are illustrated in Figure 4.13. These wetlands are
heavily influenced by regional and local groundwater patterns.

Riparian Ecosystems

Riparian ecosystems are influenced by river flood pulses, usually in the wet winter/
spring season, and dry conditions during much of the growing season. They may or
may not be jurisdictional wetlands as determined in the United States (see Chapter 15:

382 Chapter 11 Freshwater Swamps and Riparian Ecosystems

“Wetland Laws and Protection”) because of the lack of sufficient root-zone flooding
in the growing season. Riparian vegetation along a stream or river is determined by the
cross-sectional morphology, including braiding of the stream, width of the floodplain,
soil type, and elevation and moisture gradients. These are all determined in part by
larger scale (continental, basin, stream system) processes that are modified by local
biotic and physical processes. The riparian soil moisture regime, in large part, explains
the plant associations. The relationship, however, is seldom that simple. Soil moisture
and depth to groundwater are not the only factors governing plant establishment.
Low floodplain elevations are often swept clean of plants by floods, so that seedlings
do not survive, and the vegetation is limited to annuals and perennials that survive
until the next flood. Trees mature only at elevations above moderate floods where
they can become well enough established to withstand severe floods. There are distinct
differences in riparian ecosystems and floodplains in mesic and arid climates.

Mesic Riparian Ecosystems

Most of the extensive riparian ecosystems of south and eastern United States, which
include the alluvial cypress swamps described above, are characterized as zones of
deposition. These river systems are dominated by spring floods and late-summer flow
minima. A typical broad floodplain in mesic climates such as eastern North America
contains eight major features (Fig. 11.4):

1 . The river channel meanders through the area, transporting, eroding, and
depositing alluvial sediments.

2. Natural levees adjacent to the channel are composed of coarse materials that
are deposited when floods flow over the channel banks. Natural levees,
sloping sharply toward the river and more gently away from the floodplain,
are often the highest elevation on the floodplain.

3. Point bars are areas of sedimentation on the convex sides of river curves. As
sediments are deposited on the point bar, the meander curve of the river
tends to increase in radius and migrate downstream. Eventually, the point bar
begins to support vegetation that stabilizes it as part of the floodplain.

4. Meander scrolls 2SC depressions and ridges on the convex side of bends in the
river. They are formed from point bars as the stream migrates laterally across
the floodplain. This type of terrain is often referred to as ridge and swale
topography.

5. Oxbows^ oxbow lakes, or billabon^s (in Australia) are bodies of permanently
standing water that result from the cutoff of meanders. Deepwater swamps or
freshwater marshes often develop in oxbows.

6. Sloughs are areas of dead water that form in meander scrolls and along
valley walls. Deepwater swamps can also form in the permanently flooded
sloughs.

7. Backswamps are deposits of fine sediments that occur between the natural
levee and the valley wall or terrace.

Geomorphology and Hydrology 383

Figure 11.4 Major river geomerphic features of mesic riparian ecosystems (fioodplains)
including naturai levees, meander scrolls, oxbow lakes, back swamp, and river meanders
with peint bars. (Ftom Mitsch and Jorgensen, 2004)

8. Terraces are “abandoned floodplains” that may have been formed by the
river’s alluvial deposits but are not hydrologically connected to the current
river.

Two major aggradation processes are thought to be responsible for the forma-
tion of most floodplains: deposition on the inside curves of rivers (point bars) and
deposition from overbank flooding. “As a river moves laterally, sediment is deposited
within or below the level of the bankfull stage on the point bar, while at overflow
stages the sediment is deposited on both the point bar and over the adjacent flood
plain” (Leopold et al., 1964). The resulting floodplain is made up of alluvial sediments
(or alluvium) that can range from 10 to 80 m thick. Degradation (downcutting) of
floodplains occurs when the supply of sediments is decreases, a condition that could
be caused naturally with a shift in climate or with the construction of an upstream
dam. These processes are difficult to observe over short periods; both aggradation
and degradation can be inferred only from the study of floodplain stratigraphy or
long-term mapping.

384 Chapter 11 Freshwater Swamps and Riparian Ecosystems

Arid Riparian Ecosystems

Structurally, the temporal and spatial stability of these riparian ecosystems and rivers
in arid regions is fundamentally different from those of rivers in mesic climates. In
contrast to the broad, flat, expansive southeastern U.S. riparian forests, for example,
riparian ecosystems in arid regions, such as western United States, tend to be narrow,
linear features of the landscape, often lining streams with steep gradients and narrow
floodplains. Because of the dramatic differences in peak floods compared to mean
flows, the temporally unpredictable nature of flooding, and the coarseness of most
sedimentary material in this region, arid-region channels are time-dependent systems
that seldom reach any kind of equilibrium. As with mesic rivers, the two primary factors
governing channel morphology are the sediment supply and flow variability.

Biogeochemistry

Lockaby and Walbridge (1998) described the biogeochemistry of forested wedands
as “the most complex and difficult to study with any forest ecosystem type.” Forested
wedands have soil and water chemistry that varies from the rich sediments of alluvial
cypress swamps to the extremely low mineral and acidic waters of surface water depres-
sion red maple swamps and cypress domes. Wide ranges of pH, dissolved substances,
and nutrients are found in the soils and waters of these swamps. Three facts should be
noted from this wide range of soil and water chemistry:

1 . Swamps are generally acidic to circumneutral, depending on the accumulation
of peat and the degree to which precipitation dominates the hydrology.

2. Nutrient conditions vary from nutrient- and mineral-poor conditions in
rainwater-fed swamps to nutrient- and mineral-rich conditions in alluvial river
swamps and groundwater discharge swamps.

3. An alluvial river swamp often has water quality very different from that of the
adjacent river. Swamps in alluvial settings are generally fed by both
groundwater discharge and flooding rives and can have water chemistry quite
different from either source.

Many freshwater swamps, particularly alluvial river swamps, are “open” to river
flooding and other inputs of neutral and generally well-mineralized waters. The pH
of many alluvial swamps in the southeastern United States is 6 to 7, and there are
high concentrations of dissolved ions. Cypress domes and perched-basin swamps, in
contrast, are fed primarily by rainwater and have acidic waters, usually in the pH
range of 3.5 to 5.0, caused by humic acids produced within the swamp. Colloidal
humic substances contribute to both the low pH and the tea-colored or “blackwa-
ter” appearance of the standing water in many forested wedands. Isolated swamps,
such as cypress domes, have much in common with the oligotrophic or ombrotrophic
peadands described in Chapter 12: “Peadands.” Swamps open to major surface water
and groundwater inputs, however, are generally rich in alkalinity, dissolved ions, and
nutrients. For example, conducdvity of surface water ranges from only 60 pS/cm in

Vegetation 385

Table 11.1 Soil chemistry of Atlantic white cedar (Chamaecyparis thyoides) and red maple
(Acer rubrum) swamps in Maryiand compared to nonforested peatiands

Soil Parameters (top 50 cm)

White Cedar Swamp

Red Maple Swamp

Nonforested Peatiands

5.34

4.23

4.54

Organic matter (percent)

59 ± 5

67 ± 3

68 ± 2

Nitrogen (percent)

1.6 ± 0.1

1.5 ± 0.1

1.7 ± 0.1

Phosphorus (percent)

0.07 ± 0.01

0.24 ± 0.03

0.10 ± 0.01

NO3-N (pg/g)

0.8 ± 0.1

0.3 ± 0.1

0.5 ± 0.1

NH^-N (pg/g)

67 ± 4

72 ± 19

76 ± 10

Ca2+ (pg/g)

1,810

339

710

Mg2+ (pg/g)

1,420

493

477

K+ (pg/g)

1,054

1,622

857

Na+ (pg/g)

841

134

383

Fe (mg/g)

6.3

5.9

5.4

Al (mg/g)

8.0

5.4

7.6

Source’. Whigham and Richardson (1988)

cypress domes in Florida to 200 to 400 pS/cm in alluvial cypress swamps in Kentucky
and Illinois.

In a comparison of an Atlantic white cedar swamp with adjacent forested wetlands
and nonforested peatiands, Whigham and Richardson (1988) found cedar swamp
soils to be significantly higher in pH, calcium, and magnesium than the other sites
(Table 1 1. 1), suggesting a groundwater or brackish-water source might be important
for Atlantic white cedar to compete with other swamp trees. Phosphorus was lowest
in the white cedar swamp, suggesting this was the most significant limiting nutrient.
The high pH measured in this study suggests that Atlantic white cedar may do best in
sites with high pH, although these swamps have been reported to occur under low-pH
(3. 2^.4) conditions in the Great Dismal Swamp (F. Day, 1984).

In riparian forest soils, phosphorus availability has been shown to increase during
floods although the exact reason for this is often unclear. Wright et al. (200 1 ) examined
the availability of P after experimentally flooding plots in a Georgia floodplain forest
and found that flooding did release P; however they found that there was no change in
Fe/Al phosphates. The reduction of Fe'*'^ phosphates and hydrolysis of Al phosphates
has often been credited with increased P availability after soils have been flooded. While
this may occur when upland soils are flooded, the authors attributed the increased
available P during floods to biological processes such as the release of P from microbial
biomass and the suppression of biological P demand during anaerobic conditions.

Vegetation

Cypress Swamps

Southern deepwater swamps, particularly cypress wetlands, have plant communities
that either depend on or adapt to the almost continuously wet environment. There are

386 Chapter 11 Freshwater Swamps and Riparian Ecosystems

Table 11.2 Distinction between baid cypress and pond cypress swamps

Characteristic

Baid Cypress Swamp

Pond Cypress Swamp

Dominant cypress

Taxodium distichum

Taxodium distichum var. imbricarium

Dominant tupelo or

Nyssa aquatica (water tupeio)

Nyssa syivatica var. bifiora (black gum)

gum (when present)

Tree physiology

Large, oid trees, high growth rate,
usuaiiy abundance of knees
and spreading buttresses

Smaller, younger trees, low growth
rate, some knees and buttresses
but not as pronounced

Location

Aliuviai floodpiains of Coastai
Plain, particuiariy aiong Atlantic
seaboard, Guif seaboard, and
Mississippi embayment

"Uplands” of Coastal Plain,
particularly in Florida and southern
Georgia

Chemicai status

Neutrai of siightiy acid, high in
dissolved ions, usuaiiy high in
suspended sediments and rich
in nutrients

Low pFI, poorly buffered, low in
dissolved ions, poor in nutrients

Annuai fiooding from

Yes

river

Types of deepwater
swamps

Aliuviai river swamp, cypress
strand, iake-edge swamp

Cypress dome, dwarf cypress swamp

several distinctions between bald cypress and pond cypress swamps. The dominant
canopy vegetation found in alluvial river swamps of the southeastern United States
includes bald cypress {Taxodium distichum) and water tupelo (Nyssa aquatica). The
trees are often found growing in association in the same swamp, although pure stands
of either bald cypress or water tupelo are also frequent in the southeastern United
States. Many of the pure tupelo stands may have been the result of the selective logging
of bald cypress. The pond cypress-black gum {Taxodium distichum imbricarium
[Nutt. sylvatica var. biflora [Walt.] Sarg.) swamp is more commonly found
on the uplands of the southeastern Coastal Plain, usually in areas of poor sandy soils
without alluvial flooding (Table 11.2). These same conditions are usually found in
cypress domes.

One of the main features that distinguishes bald cypress trees from pond cypress
trees is the leaf structure (Fig. 11.5). Bald cypress has needles that spread from the
twig in a flat plane, whereas pond cypress needles are appressed to the twig. Both
species are intolerant of salt and are found only in freshwater areas. Pond cypress is
limited to sites that are poor in nutrients and are relatively isolated from the effects of
river flooding or large inflows of nutrients.

When deepwater swamps are drained or when their dry period is extended
dramatically, they can be invaded by pine (e.g., Pinus dliottii) or hardwood species.
In north-central Florida, a cypress-pine association indicates a drained cypress dome
(Mitsch and Ewel, 1979). Hardwoods that characteristically are found in cypress
domes include swamp red bay {Persea palustris) and sweet bay {Majjnolia virjjiniana) .
In lake-edge and alluvial river swamps, several species of ash {Fraxinus sp.) and maple
{Acer sp.) often grow as subdominants with the cypress or tupelo or both. In the

Vegetation 387

Pond cypress
Taxodium distichum var, imbricarium

Bald cypress
Taxodium distichum

Figure 11.5 Distinction of leaves between top: bald cypress (Taxodium distichum) and bot-
tom: pond cypress (Taxodium distichum var. imbricarium; formerly known as T. distichum var.
nutans).

Deep South, Spanish moss ( Tillandsia usneoides) is found in abundance as an epiphyte
on the stems and branches of the canopy trees.

The abundance of understory vegetation in cypress-tupelo swamps depends on
the amount of light that penetrates the tree canopy. Many mature swamps appear as
quiet, dark cathedrals of tree trunks devoid of any understory vegetation. Even when
enough light is available for understory vegetation, it is difficult to generalize about its
composition. There can be a dominance of woody shrubs, of herbaceous vegetation, or
of both. Fetterbush {Lyonia lucida)., wax myrtle {Myrica cerifera), and Virginia willow
{Itea virjjinica) are common as shrubs and small trees in nutrient -poor cypress domes.
Understory species in higher-nutrient river swamps include buttonbush ( Cephalanthus

388 Chapter 11 Freshwater Swamps and Riparian Ecosystems

occidentalis) and Virginia willow. Some continually flooded cypress swamps that have
high concentrations of dissolved nutrients in the water develop dense mats of duck-
weed (e.g., Lemna spp., or Spirodda spp., or Azolla spp.) on the water surface during
most of the year. Floating logs and old tree stumps often provide substrate for under-
story vegetation to attach and to flourish.

White Cedar Swamps

Cedar swamps occur withm a wide climatic range along the East Coast of the United
States and in an intermediate hydrology between deepwater cypress swamps in the
South and forested swamps such as red maple swamps in the North. Often these
swamps are monospeciflc, even-aged stands with tightly spaced Chamaecyparis thy-
oidrr trees, and no subcanopy, few shrubs, and minimal herbaceous plants. However,
the tree is often found in mixed stands, with co-dominants such as Betula popuUfolm
(gray birch), Picea mariana (black spruce), Finns strobus (Eastern white pine), and
Tsu^a canadensis (Eastern hemlock) (Laderman, 1989). In the South co-dominant
trees include Gordonia lasianthus (loblolly bay), Persea borbonia (red bay), P. palustris
(swamp red bay), and Taxodium districhum (bald cypress).

The shrub layer in cedar swamps with relatively open canopies includes many eri-
caceous shrubs, such as Aronia arbutifolia (red chokeberry), Clethra alnifoUa (sweet
pepperbush), Ilexpilabra (gallberry), Leucothoe racemosa (fetterbush), and Vaccinium
corymbosum (highbush blueberry) (Laderman, 1989).

Red Maple Swamps

The canopy of red maple swamps is obviously dominated by Acer rubrum L. Canopy
cover generally exceeds 80 percent, although trees in these northern swamps tend to
be shorter with less biomass than those in southern swamps. Although up to 50 tree
species have been found in a red maple swamp, the red maple can account for up
to 90 percent of the stem density and basal area (Golet et al., 1993). In general, a
speciflc site will have about four species of trees in the canopy/subcanopy, depending
on which region of the glaciated Northeast these red maple swamps occur.

Shrubs include Ilex vertucillata (winterberry), Vaccinium corymbosum (highbush
blueberry), Lindera benzoin (spicebush). Viburnum spp. (arrowwood), Alnus ru^osa
(speckled alder), Cephalanthus occidentalis {huttonhwsh), Corylus cornuta (hazelnut),
and Rhododendron viscosum (swamp azalea), with dominance depending on the region
in which the swamps are found. Shrub cover is generally greater than 50 percent,
although some red maple swamps have shrub cover as low as 6 percent. One of
the most interesting features of many red maple swamps is the predominance of a
great variety of ferns in the herbaceous layer, including Osmunda cinnamomea (cinna-
mon fern), Onoclea sensibilis (sensitive fern), Osmunda regalis (royal fern), Thelypteris
thelypteroides (marsh fern), Matteuccia struthiopteris (ostrich fern), Osmunda clayto-
niana (interrupted fern), and various Dryopteris spp. (wood ferns). Other common
herbaceous plants include Symplocarpus foetidus (skunk cabbage), Caltha palustris

Vegetation 389

(marsh marigold), several species of Glyccrm (manna grass), and several of more than
32 species of Carex.

Riparian Ecosystems

Southeastern U.S. Bottomland Forests

The vegetation of high-order southeastern riparian ecosystems is dominated by
diverse trees that are adapted to the wide variety of environmental conditions on the
floodplain. The most important local environmental condition is the hydroperiod,
which determines the “moisture gradient,” or — as Wharton et al. (1982) prefer — the
“anaerobic gradient,” which varies in time and space across the floodplain. The plant
species found along this gradient respond to elevation relative to the river’s flooding
regime (Fig. 11.6). The lowest parts of the bottomland, nearly always flooded, form
cypress-tupelo gum swamps above. At slightly higher bottomland elevations than
the deep swamps, the soils are semipermanendy inundated or saturated and support
an association of black willow (Salix silver maple [Acer saccharinum), and

sometimes cottonwood {Populus deltoides) in the pioneer stage. A more common

Channel Levee

Upland Forest

Flooding

extent

Flooding
frequency.
% of years

Flooding

duration,

% of growing
season

Tree

^)ec»es

Submorged
aquatics
open water

Seasonalty

mtenninently

Semi-

lemporanly

flooded

11 50

51 - 100

-100

2*25

>25

-100

Sycamore

Overcup oak

Bald cypress

hatanus

Ouercus tyraia

Taxodium

ooctdentaks

Water mduxy

dtsdctnjm

Sweetgum

Carya aquatica

Water tupeio

Uqtatian^uu

Green ash

Nyssa

styraaHua

Ftaxinus

aquatKa

Amencan elm

pennsyfvanica

Red maple

UfrTNi5

Sugarberry

Acer rubrum

americana

Ceths iaevigata

Green asf)

Amencan elm

Fiaxinus

Vkrms

amaficana

pannsyh/amca

Sweetgum

LiquKiat^^r StyfsctlUia
Willow oak
OuefCJsphe^io&
Water oak
OuewMS mgra
cr>erT)i>ar1( oak
O. toJeatA var poQOdttoka
Swamp chestmil oak
Q. michauKu

Seldom

flooded

upland lorest
Species

Figure 11.6 General relationship between vegetation associations and floodplain topo-
graphy, flood frequency, and flood duration of a southeastern United States bottomland
hardwood forest. (Ftom Mitsch and Gossslink, 2000)

390 Chapter 11 Freshwater Swamps and Riparian Ecosystems

association in this zone includes overcup oak {Quercus lymta) and water hickory
{Carya aquatica), which often occur in relatively small depressions on floodplains.
Also tbund in this zone are green ash {Fraxinus pennsylvanica), red maple {Acer
rubrum), and river birch {Betula nipira). Higher still on the bottomland floodplain in
areas flooded or saturated one to two months during the growing season are found an
even wider array of hardwood trees, including laurel oak {Quercus laurifoUa)^ green
ash {Fraxinus pennsylvanica)^ American elm ( Ulmus americana), and sweetgum
{Liquidambar styraciflua) as well as sugarberry {Celtis laevigata)^ red maple {Quercus
rubra)^ willow oak {Quercus phellos)^ and sycamore {Platanus occidentalis) . Pioneer
successional communities in this zone can consist of monotypic stands of river birch
or cottonwood.

Temporarily or infrequendy flooded terraces at the highest elevations of the flood-
plain (second terrace in Fig. 11.6) are flooded for less than a week to about a month
during each growing season and are often dominated by several oaks, tolerant of
occasionally wet soils such as swamp chestnut oak {Quercus michauxii)^ cherrybark
oak {Quercus falcata var. pagodifoUa)^ and water oak {Quercus n{0ra) and hickories
{Carya spp.).

Plant zonation is not linear topographically, nor is it vegetationally discrete.
Figure 11.6 is a cross section of the micro topography of an alluvial floodplain in
the southeastern United States. In reality, the complex microrelief does not show
a smooth change from one zone to the next. The natural levee next to the stream
(Fig. 11.6), in fact, is often one of the most diverse parts of the floodplain because of
fluctuations in its elevation.

Arid and Semiarid Riparian Forests

The vegetation in riparian forests of the semiarid grasslands and arid western United
States differ from those found in the humid eastern and southern United States. The
natural upland ecosystems of this region are grasslands, deserts, or other nonfbrested
ecosystems, and so the riparian zone is a conspicuous feature of the landscape. West-
ern U.S. riparian ecosystem tree species are phreatophytes\ that is, they are plants that
obtain their water from phreatic sources (i.e., groundwater or the capillary fringe of
the groundwater table). Many species use surface water supplies when seedlings (hence
the general germination requirement of bare, moist soil) but put down long, deep
roots that later supply water requirements from groundwater. Cottonwoods {Populus
spp.) are considered obligate phreatophytes, while both Prosopis pubescens {mcsqaitc)
and alien Tamarix ramisissima (salt cedar) are facultative. Salt cedar is an introduced
species that is rapidly replacing cottonwood in many areas.

Swamps of Glaciated Regions

Forested swamps occur throughout the glaciated midwestern United States; in fact,
most of the wetlands remaining in states such as Ohio, Indiana, and Illinois are
forested wedands that occur in isolated basins or floodplains amid agricultural Adds
(Table 1 1.3). They were the flelds that were too wet to plant and gradually were

Vegetation 391

Table 11.3 Typical vegetation in a hardwood swamp forest in centrai Ohio^

Trees

Wetland indicator Status'’

Trees

Quercus palustrus (pin oak)

FACW

Quercus bicolor (swamp white oak)

FACW

Acer saccharinum (silver maple)

FACW

Acer rubrum (red maple)

FAC

Ulmus americana (American elm)

FACW

Fraxinus pennsylvanica (green ash)

FACW

Shrubs/Understory

Lindera benzoin (spicebush)

FACW

Cephalanthus occidentalis (buttonbush)

OBL

Rosa multiflora^ (multiflora rose)

FACU

Carpinus caroliniana (hornbeam, ironwood)

FAC

Herbs

Polygonum spp. (smartweed)

FAC/OBL

Symplocarpus foetidus (skunk cabbage)

OBL

Lemna spp. (duckweed)

OBL

Alisma plantago-aquatica (water plantain)

OBL

Aster spp. (asters)

FAC/ FACW

Carex spp. (sedges)

FACW/OBL

Ranunculus septentrionalis (swamp buttercup)

OBL

Saxifraga pennsylvanica (swamp saxifrage)

OBL

Onoclea sensibilis (sensitive fern)

FACW

Bidens comosa (leafy-bracted beggar-ticks)

FACW

Bidens frondosa (devil's beggar-ticks)

FACW

Scirpus atrovirens (green bulrush)

OBL

Scirpus cyperinus (wool grass)

FACW

^Wetland species at Gahanna Woods Nature Preserve, Franklin County, Ohio.

'’Use of wetland indicator status for the northeastern United States. In order of wet to
dry: OBL = obligate wetland plant; FACW = facultative wet plant; FAC = facultative plant;
FACU = facultative upland plant.

‘’Nonnative species.

invaded by tree species. They often are a remnant of a gradual process of ponds of
glacial origin slowly infilling and becoming forested (a true hydrarch succession).
However, they may also occur in wet basins on mineral hydric soils rather than peat
deposits. As with red maple swamps, the trees generally have replaced herbaceous
marshes that once occupied those sites, because of natural succession or because of
artificial drainage. The succession of these systems is poorly understood.

Tree Adaptations

Vascular plants, particularly trees, have a difficult time surviving under continuously
flooded conditions. Only a handful of species of trees in North America can stay viable
in continuous flooding, and, even then, their growth is generally slowed; trees that

392 Chapter 11 Freshwater Swamps and Riparian Ecosystems

Fires in Swamps?

Fire is generally infrequent in swamps because of standing water or saturated
soil conditions, but It can be a significant ecological factor during droughts or
in swamps that have been artificially drained. In general, fire is more frequent
in the forested swamps of Florida than anywhere else, because of the more fre-
quent lightning storms and because of a predictable dry season. For example,
from 1970 to 1977, there were four fires in the Big Cypress National Preserve
in southern Florida, each affecting an average of 500 ha. In April-May 2009,
a fire burned a much larger area (12,000 ha) of the northwestern part of the
preserve (Watts et al., 2012). Fire is rare in most alluvial river swamps but can
be more frequent in cypress domes or dwarf cypress swamps — as frequent as
several times per century.

Fire had a “cleansing” effect on the trees in a cypress dome in
north-central Florida in the sense that the fire selectively killed almost all of
the upland pine and hardwoods that had invaded the cypress dome but left
the cypress unharmed (Ewel and Mitsch, 1978). This suggests a possible
advantage of fire to some shallow cypress ecosystems in eliminating compe-
tition that is less water tolerant. Casey and Ewel (2006) identified fire severity
as a key factor influencing tree succession in Florida pond cypress swamps. In
their generalized succession model, the exclusion of fire (due to geomorphic
conditions) tends to promote mixed bay-cypress communities, while periodic
moderate fires tend to promote monotypic cypress or cypress-tupelo forests.
Severe fires can lead to shrub or marsh conditions.

Fire can also be an influential factor on white cedar swamps. If water is
low, fire can be quite destructive, killing cedar trees and burning the peat
deeply. If water levels are high, light fire can have a cleansing effect, elimi-
nating shrubs and brush and favoring cedar seedling germination (Laderman,
1989). In C. thyoides swamps of the Atlantic Coast, the highly flammable cedar
foliage burned frequently (five fires per each 100- to 200-year interval) during
pre-European settlement time; when fires became more rare after European
settlement, stands of cedar became the familiar dense monospecific systems
that are common today (Motzkin et al., 1993).

Watts (2013) and Watts and Kobziar (2013) described fires in wetlands
such as cypress swamps as often being smo/der/ng combustion or ground fires,
as opposed to flaming combustion typical of fires in upland forests. These fires
can continue for many days or even months, are much more difficult to control
than are flaming fires, and produce an additional human hazard of abundant
smoke, day and night.

are found in freshwater swamps are stressed with the wet conditions but have found
ways to adapt. The most conspicuous adaptations specifrc to the major tree species in
forested swamps are discussed here.

Vegetation 393

Knees and Pneumatophores

Cypress (bald and pond), water tupelo, and black gum are among a number ofwetland
plants that produce pneumatophores. In deepwater swamps, these organs extend from
the root system to well above the average water level (Fig. 11.7a). On cypress, these
“knees” are conical and typically less than 1 m in height, although some cypress knees
are as tall as 3 to 4 m. Knees are generally much more prominent on cypress than
on tupelo. Pneumatophores on black gum in cypress domes are actually arching or
“kinked” roots that approximate the appearance of cypress knees. The functions of the
knees have been speculated about for more than a century. It was thought that knees
might be adaptations for anchoring trees because of the appearance of a secondary root
system beneath knees that is similar to and smaller than the trees’ main root system.
Observations of swamp and upland damage in South Carolina following Hurricane
Hugo in 1990 showed that cypress trees often remained standing while hardwoods
and pines did not, supporting the tree -anchoring theory for cypress root, knee, and
buttress systems (K. Ewel, personal communication).

(c)

Figure 11.7 Among the several features of vegetation in cypress swamps are (a) cypress
knees, (b) cypress tree butresses, (c) large size and long life of cypress trees.

394 Chapter 11 Freshwater Swamps and Riparian Ecosystems

Other discussions of cypress knee function have centered on their possible use
as sites of gas exchange for the root systems. Penfound (1952) argued that cypress
knees are often absent where they are most needed — in deep water — and that the
wood of the cypress knee is not aerenchymous; that is, there are no intercellular gas
spaces capable of transporting oxygen to the root system. However, gas exchange
does occur at the knees. S. L. Brown (1981) estimated that gas evolution from knees
accounted for 0.04 to 0.12 g C m~^ yr“^of the respiration in a cypress dome, and
0.23 g C m”^ yr~^^ in an alluvial river swamp. This accounted for 0.3 to 0.9 per-
cent of the total tree respiration but 5 to 15 percent of the estimated woody tissue
(stems and knees) respiration. The fact that carbon dioxide (CO2) is exchanged at the
knee, however, does not prove that oxygen transport is taking place there or that the
CO2 was the result of oxidation of anaerobically produced organic compounds in the
root system.

Buttresses

Taxodium and Nyssa species and, to a lesser degree, Chamaecyparis thyoides often
produce swollen bases or buttresses (stem hypertrophy) when they grow in flooded
conditions (Fig. 11.7b). The basal swelling can extend from less than 1 m above the
soil to several meters, depending on the hydroperiod of the wetland. Swelling gener-
ally occurs along the part of the tree that is flooded at least seasonally, although the
duration and frequency of the flooding necessary to cause the swelling are unknown.
One theory described the height of the buttress as a response to aeration: The greatest
swelling occurs where there is a continual wetting and soaking of the tree trunk but
where the trunk is also above the normal water level (Kurz and Demaree, 1934). The
value of the buttress swelling to ecosystem survivability is unknown; it may simply be
a relict response that is of littie use to the plant.

Seed Germination and Dispersai

The seeds of swamp trees require oxygen for germination. For example, cypress seeds
and seedlings require moist but not flooded soil for germination and survival. Occa-
sional drawdowns, if only at relatively infrequent intervals, are therefore necessary for
the survival of trees in these swamps unless floating mats develop. Otherwise, contin-
uous flooding will ultimately lead to an open-water pond.

The dispersal and survival of the seeds of many swamp trees depend on hydro-
logic conditions. Schneider and Sharitz ( 1986) found a relatively low number of viable
seeds in a seed bank study of a cypress-tupelo swamp in South Carolina. An aver-
age of 127 seeds/m^ were found for woody species (88 percent cypress or tupelo)
in the swamp compared to a seed density of 233 seeds/m^ from an adjacent bot-
tomland hardwood forest. The authors speculated that the continual flooding in the
cypress-tupelo swamp leads to reduced seed viability. Huenneke and Sharitz (1986)
elaborated further on the importance of hydrochory (seed dispersal by water) in these
swamps. Hydrologic conditions, particularly scouring by flooding waters, are impor-
tant factors in determining the composition, dispersal, and survival of seeds in riverine
settings. Seeds are transported relatively long distances; the highest seed densities

Consumers 395

accumulate near obstructions such as logs, tree stumps, cypress knees, and tree stems,
and the lowest seed densities occur in open-water areas.

Longevity

Some swamp trees may live for centuries and achieve great sizes (Fig. 11.7c). One
individual bald cypress tree in Corkscrew Swamp in southwestern Florida was deter-
mined to be about 700 years old. Laderman (1998) reported that the maximum
age of Taxodium is 1,000 years. By contrast, Chcmmecyparis thyoides lives to a max-
imum of 300 years (Clewell and Ward, 1987). Mature bald cypress trees are typi-
cally 30 to 40 m in height and 1 to 1.5 m in diameter. Anderson and White (1970)
reported a very large cypress tree in a cypress-tupelo swamp in southern Illinois that
measured 2.1m in diameter. C. A. Brown (1984) summarized several reports that
documented bald cypress as large as 3.6 to 5.1 m in diameter.

Shallow or Adventitious Roots

Some species, such as Acer rubrum, develop very shallow root systems in response
to flooding, in all likelihood because the surface soil is closest to the atmospheric
source of oxygen. In aerated soils, the same species will develop deep roots. Other
swamp species, such as willows [Salix sp.), green ash {Fraxinus pennsylvcmica)^ and
cottonwoods {Populus deltoides), develop adventitious roots above-ground from the
stem in response to flooding.

Gaseous Diffusion

Woody trees have a particular problem getting oxygen to their rhizosphere when they
are flooded, and few species do it well enough to survive continual flooding. The
swamp trees, including Taxodium^ Nyssa, Alnus, and Fraxinus, among others, have
the ability to supply oxygen to their root systems in amounts adequate for rhizospheric
demands. Solar radiation, which heats up the tree stems by a couple of degrees, causes
a light-induced gas flow that can be considerably greater in selected swamp seedlings
than in the same trees in the dark: This thermally induced flow of air through vas-
cular plants is called thermo -osmosis by some (Grosse et ah, 1998) and is enhanced
by the development of aerenchymous stem and root tissue (see Chapter 7: “Wetland
Vegetation and Succession”).

Consumers

Invertebrates

Invertebrate communities, particularly benthic macroinvertebrates, have been ana-
lyzed in several cypress-tupelo swamps. A wide diversity and high number of inver-
tebrates have been found in permanently flooded swamps. Species include crayfish,
clams, oligochaete worms, snails, freshwater shrimp, midges, amphipods, and vari-
ous immature insects. Batzer and Wissinger (1996) reported that insects, particularly
midges, can dominate forested wetlands and that midges are most likely to reach high

396 Chapter 11 Freshwater Swamps and Riparian Ecosystems

densities. Many of these invertebrates are highly dependent, either directly or indi-
rectly, on the abundant detritus found in these systems.

Oligochaetes and midges (Chironomidae), both of which can tolerate low-
dissolved-oxygen conditions, and amphipods such as Hyalella azteca^ which occur
in abundance amid aquatic plants such as duckweed, usually dominate alluvial river
swamp invertebrate communities. In nutrient-poor cypress domes, the benthic fauna
are dominated by Chironomidae, although crayfish, isopods, and other Diptera
are also found there. Stresses stemming from low dissolved oxygen and periodic
drawdowns account for the low diversity and number in these domes.

The production of wood in deepwater swamps results in an abundance of
substrate for invertebrates to colonize, although few studies have documented the
importance of this substrate in swamps tor invertebrates. Thorp et al. (1985) found
that suspended Nyssa logs had three times as many invertebrates and twice as many
taxa when they were placed in a swamp-influent stream than in the swamp itself
or by its outflow stream. The swamp inflow had the highest number of mayflies
(Ephemeroptera), stoneflies (Plecoptera), midges (Chironomids), and caddie flies
(Trichoptera), whereas Oligochaetes were greatest in the swamp itself, supposedly
because of anoxic, stagnant conditions. Understory plants within swamps have also
been shown to be important to various invertebrate groups.

Fish

Fish are both temporary and permanent residents of alluvial river swamps. Several stud-
ies have noted the value of sloughs and backswamps for fish and shellfish spawning and
feeding during the flooding season. Forested swamps often serves as a reservoir for fish
when flooding ceases, although the backwaters are less than optimum for aquatic life
because of fluctuating water levels and occasional low-dissolved-oxygen levels. Some
fish such as bowfin {Amia calva), gar {Lepisosteus sp.), and certain top minnows (e.g.,
Fundulus spp. and Gambusm affinis) are better adapted to periodic anoxia through
their ability to utilize atmospheric oxygen. Several species of forage minnows often
dominate alluvial river swamps, where larger fish are temporary residents of the wet-
lands. Fish are sparse to nonexistent in the shallow cypress domes, white cedar swamps,
and red maple swamps because of the lack of continuous standing water.

Reptiles and Amphibians

Reptiles and amphibians are prevalent in swamps because of their ability to adapt to
fluctuating water levels. Nine or 10 species of frogs are common in many southeastern
cypress-gum swamps. Two of the most interesting reptiles in southeastern deepwater
swamps are the American alligator {Alligator mississippimsis) and the cottonmouth
moccasin {A^kistrodon piscivorus) . The alligator ranges from North Carolina through
Fouisiana, where alluvial cypress swamps and cypress strands often serve as suitable
habitats. The cottonmouth, or water moccasin, a poisonous water snake that has a
white inner mouth, is found throughout much of the range of cypress wetlands and is

Ecosystem Function 397

the topic of many a “snake story” of those who have been in these swamps. Other water
snakes, particularly several species of Nerodia, however, are often more important in
terms of number and biomass and often are mistakenly identified as cottonmouth.
The snakes feed primarily on frogs, small fish, salamanders, and crayfish.

Red maple swamps are important areas in the forested northeastern United States
for the breeding and feeding of reptiles and amphibians. DeGraaf and Rudis (1986)
found that 45 species of reptiles and amphibians required forest cover sometime during
the year in New England and that of the 1 1 types of forests studied, red maple swamps
were actually the preferred habitat of 12 of those 45 species. In a later study, DeGraaf
and Rudis (1990) found that red maple swamps with streams supported twice as many
individuals of reptiles and amphibians as did red maple swamps without streams, with
wood frog {Rana sylvatica)^ redback salamander {Plethodon cinereus)^ and American
toad {Bufo americanus) accounting for 90 percent of the abundance.

Ecosystem Function

Four generalizations about the ecosystem function of freshwater swamps will be dis-
cussed in this section:

1. Swamp productivity is closely tied to its hydrologic regime.

2. Nutrient inflows, often coupled with hydrologic conditions, are major
sources of influence on swamp productivity.

3. Swamps can be nutrient sinks whether the nutrients are a natural source or
are artificially applied.

4. Decomposition of woody and nonwoody material in swamps is affected by
the water regime and the subsequent degree of anaerobiosis.

Primary Productivity

The importance of flood pulsing (the flood stability concept of W. E. Odum et al.,
1995) to the productivity of swamps is illustrated in Figure 11.8a, where the basal-area
growth of bald cypress in an alluvial river swamp in southern Illinois was strongly
correlated with the annual discharge of the adjacent river. This graph suggests that
higher tree productivity in this wetland occurred in years when the swamp was flooded
more frequently than average or for longer durations by the nutrient-rich river. Similar
correlations were also obtained when other independent variables that indicate degree
of flooding were used.

The importance of nutrient inflows as well as hydrologic conditions to produc-
tivity in cypress swamps in general is illustrated in Fig. 11.8b. Hydrologic inflows and
nutrient inflows are coupled in most swamps, so both charts in Figure 11.8 reflect
the same phenomenon. There is a wide range of productivity reported for forested
swamps, with almost all of the studies carried out in the southeastern United States
(Table 11.4). Primary productivity depends on hydrologic and nutrient conditions

Figure 11.8 Relationships between hydroiogic conditions and tree productivity in cypress
swamps: (a) increase in basal area of bald cypress trees in southern liiinois alluvial swamp
as a function of river discharge for five-year periods, and (b) biomass production as a func-
tion of phosphorus infiow for several cypress swamps. Data points in (a) indicate mean; bar
indicates 1 standard error, ((a) After Mitsch et al., 1979; (b) after S. L. Brown, 1981)

398

Table 11.4 Biomass and net primary productivity of deepwater swamps in the southeastern United States

■a

s t

> t

o E
< 39

•qO \
c til

<0 C53

O O
vH tH CO

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400

Ecosystem Function 401

and pulsing hydrology supports more productive systems than does permanent flood-
ing or lack of flooding. Several other studies have reported the importance of flooding
to forests by linking annual tree growth with flood occurrence (Conner and Day 1976,
Robertson et al. 2001, Stromberg, 2001, Anderson and Mitsch, 2008).

Based on the subsidy-stress model (E. R Odum et ah, 1979), floodplain tree
growth should be maximized where flooding is frequent or long enough to subsidize
nutrients and enhancing growing conditions but not so much that floods become a
physiological stress to trees. Attempts to demonstrate this model by comparing forest
communities along a wetness gradient have often been inconclusive. Megonigal et al.
(1997) investigated productivity in floodplain swamps throughout the southeastern
United States and concluded that while permanently flooded floodplain swamps did
have lower productivity, there was no evidence that sites that were seasonally pulsed
were any more productive than sites that were clearly upland (Fig. 11.9). They sug-
gested that the Mitsch and Rust (1984) model (see Fig. 4.18 in Chapter 4) may be a
more appropriate description of the productivity of forested wedands.

In almost all of these studies, only above-ground productivity was estimated.
Powell and Day (1991) made direct measurements of below- ground productivity and
found that it was highest in a mixed hardwood swamp (989 g m~^ yr~^) and much
lower in a more frequently flooded cedar swamp (366 g yt~^), a cypress swamp
(308 g m“^ yr~^)i and a maple-gum swamp (59 g m~^ yr~^)- These results suggest
that the allocation of carbon to the root system decreases with increased flooding.

Mean growing-season water depth, cm

Figure 11.9 The relationship between net primary productivity of fioodplain forests and
mean growing season water depth in bottomland hardwood forests of the southeastern
United States. (After Megonigal et al., 1997)

402 Chapter 11 Freshwater Swamps and Riparian Ecosystems

In forested wetlands that have been unaltered, annual mortality rates of trees
are often low. Conner et al. (2002) monitored annual changes in forested wetland
structure between 1987 and 1999 in South Carolina and Louisiana. Tree mortality in
unaltered areas was low (~2 percent); however, higher annual mortality (up to 16 per-
cent) was observed at Louisiana sites where severe water-level rise has occurred. The
authors also found that severe windstorms increased short-term mortality, but these
events can also lead to elevated long-term mortality rates as damaged trees eventually
succumb.

Energy Flow

The energy flow of deepwater swamps is dominated by primary productivity
of the canopy trees. Energy consumption is accomplished primarily by detrital
decomposition. Significant differences exist, however, between the energy flow
patterns in low-nutrient swamps, such as dwarf cypress swamps and cypress domes,
and high-nutrient swamps, such as alluvial cypress swamps (Table 11.5). All of the
cypress wetlands are autotrophic — ^productivity exceeds respiration. Gross primary
productivity, net primary productivity, and net ecosystem productivity are highest in
the alluvial river swamp that receives high-nutrient inflows. Buildup and/or export
of organic matter are characteristic of all of these deepwater swamps but are most
characteristic of alluvial swamps. There are few allochthonous inputs of energy to
the low-nutrient wetlands, and energy flow at the primary producer level is relatively
low. The alluvial cypress-tupelo swamp depends more on allochthonous inputs of
nutrients and energy, particularly from runoff and river flooding. In alluvial deepwater
swamps, productivity of aquatic plants is often high, whereas aquatic productivity in
cypress domes is usually low.

Nutrient Budgets

The functioning of forested wetlands as nutrient sinks was first suggested by Kitchens
et al. (1975) in a preliminary winter-spring survey of an alluvial river swamp complex
in South Carolina. They found a significant reduction in phosphorus as the waters

Table 11.5 Estimated energy flow (kcal m~^ day~^) in seiected Florida cypress swamps^

Parameter

Dwarf Cypress Swamp

Cypress Dome

Alluvial River Swamp

Gross primary productivity*’

115

233

Plant respiration”

205

Net primary productivity

Soil or water respiration

Net ecosystem productivity

^Assume Ig C = 10 kcal.

'’Assumes gross primary productivity (GPP) = net daytime photosynthesis + nighttime ieaf respiration.
“’Piant respiration = 2 x (nighttime ieaf respiration) + stem respiration + knee respiration.

Source: S. L. Brown (1981)

Ecosystem Function 403

passed over the swamp and assumed this to be the result of biological uptake by
aquatic plant communities. In a similar study in Louisiana, J. W. Day et al. (1977)
found that nitrogen was reduced by 48 percent and phosphorus decreased by 45 per-
cent as water passed through a lake-swamp complex of Barataria Bay to the lower
estuary. They attributed this decrease in nutrients to sediment interactions, includ-
ing nitrate storage/denitrification and phosphorus adsorption to the clay sediments.
Beginning in 1973, H.T. Odum et al. (1977) and colleages and students investigated
recycling of treated sewage applied to cypress domes and other swamps in northcentral
Florida. Much of that work was later summarized in Ewel and Odum (1984). Since
then, countless studies have illustrated the potential of forested wetlands for nutrient
removal, including several studies in Louisiana (Mitsch and Day, 2004; Day et ah,
2004; Rivera-Monroy et ah, 2013).

Nutrient budgets of deepwater swamps vary from “open” alluvial river swamps
that receive and export large quantities of materials to “closed” cypress domes that
are mostly isolated from their surroundings (Table 11.6). Mitsch et al. (1979) devel-
oped a nutrient budget for an alluvial river swamp in southern Illinois and found
that 10 times more phosphorus was deposited with sediments during river flooding
(3.6 g-P m”^ yr”^) than was returned from the swamp to the river during the rest of
the year (see Fig. 6.18 in Chapter 6, “Wedand Biogeochemistry”). The swamp was a
sink for a significant amount of phosphorus and sediments during that particular year
of flooding, although the percentage of retention was low (3M:.5 percent) because a
very large volume of water passed over the swamp during flooding conditions. Noe and
Hupp (2005) evaluated net nutrient accumulation in floodplain forests along rivers
contributing to the Chesapeake Bay. Mean accumulation rates for C ranged from 61
to 212 g-C m“^ yr”^, N ranged from 3.5 to 13.4 g-N m“^ yr“\ and P ranged from 0.2
to 4.1 g-P yr“h Watershed land use was a significant factor in their study. The
greatest accumulation of sediment and nutrients occurred along the Chickahominy
River, downstream from the urban metropolitan area of Richmond, Virginia.

Table 11.6 Phosphorus inputs to forested swamps (g-P m~^ yr~^)

Sediments from

Swamp Rainfall Surface Inflow River Flooding Reference

Florida

Dwarf cypress

0.11

S. L. Brown (1981)

Cypress dome

0.09

0.12

Alluvial river swamp

—

3.1

S. L. Brown (1981)

Southern iiiinois

Alluvial river swamp

0.11

0.1

3.6

Mitsch et al. (1979)

North Carolina

Alluvial tupelo swamp

0.02-0.04

0.01-1.2

0.2

Yarbro (1983)

Virginia

Floodplain forests

0.2-4.1

Noe and Hupp (2005)

404 Chapter 11 Freshwater Swamps and Riparian Ecosystems

Riparian Ecosystems and River Exchanges

Ecologists have reviewed river systems in terms of their ecological function and have
developed two different ways of describing flowing water systems. The river contin-
uum system clearly is related to the general differences in ecology along streams and
rivers, going longitudinally along the river itself. The concepts were developed mostly
in low-order streams in the United States. Little attention is paid to lateral connec-
tions or two floodplains. The flood pulse concept, however, based on research done
in the Amazon River and its tributaries, features the importance of seasonal patterns
of stream flow and the importance of lateral exchange between the river and its ripar-
ian ecosystems.

River Continuum Concept

The river continuum concept (RCC) is a theory developed in the early 1980s to
describe the longitudinal patterns of biota found in streams and rivers (Vannote et ah,
1980; Minshall et al., 1983, 1985). According to the RCC, most organic matter
is introduced to streams from terrestrial sources in headwater areas (Fig. 11.10).
The production/respiration (P/R) ratio is < 1 (i.e., the stream is heterotrophic),
and invertebrate shredders and collectors dominate the fauna. Biodiversity is limited
by low temperatures, low light, and low nutrients. In river midreaches, more light

Dominant Energy Organic Benthic

Sources Matter invertebrates

Shredders
and Coliectors

Grazers and
Coilectors

Collectors

Figure 11.10 The river continuum cencept showing transition from smail first-order stream
to very iarge eleventh-order river. Charts on the left indicate relative importance of terres-
trial, in-stream, or upstream energy seurces to the aquatic food chain. Charts on the right
indicate the reiative importance of different feeding groups of invertebrates. (From Mitsch
and Jorgensen, 2004, after Johnsen et al., 1995)

Ecosystem Function 405

Fish Activity

Nutrient Rux

Aquatic/terrestrial transition zone

Most river-
spawning fish
start to breed

Lake and nver
spawning;
young-o^year
and predators
follow moving
littoral; fish
and Invertebrate
production high

Young and adult
fish cfTsperse and
feed, dissc^ed
oxygen (DO)
permitting

Many fish
respond to
drawdown by
finding deeper
water

Fish migrate
to main channei,
permanent lakes
or tributanes

Maximum productvity
of aquatic vegetatiori:
decomposition of
terrestrial vegetation

maximum Hooding

— ****tftj

low water

Mlneralizatbn of
nutrients from
decomposition

Concentration
of nutrients in
riyer; consolidation
of sediments

Maximum biomass
of emergent
macrophytes; high
HjS and low DO
in river bottom water

Moist soil plants
germinate; continued
decomposition of
aquatic plants

Regrowth of
terrestrial grasses
and shrubs

Figure 11.11 The fleod pulse cencept for a river and its fleodplain, illustrating five periods
over the wet and dry seasons of a river. (From Mitsch and Jorgensen, after Bayley, 1995, and
Junk et al., 1989)

is available, phytoplankton prospers, and biodiversity is highest. The P/R ratio is
>1. Organic matter input from upstream is fine; filter feeders dominate the flora. In
braided reaches or where the floodplain is broad, however, the bank habitat is a major
source of snags and logs that lead to debris dams that slow water flow and increase
stream habitat diversity. The increased input of riparian coarse debris increases food
diversity and increases heterotrophy. The productivity/respiration (P/R) ratio is <1.
Finally, in the highest-order streams, riparian litter inputs are minor and turbidity
reduces primary productivity. Hence the system is heterotrophic again (P/R <1),
and diversity is often low. The importance of backwaters, oxbows, and floodplains to
river ecosystem function are virtually ignored in the RCC.

Flood Pulse Concept

The RCC considers the importance of the riparian zone only in an indirect way by not-
ing that small low-order streams are influenced by shading and abundant contributions
of allochthonous organic matter. Junk et al. (1989) developed z. flood pulse concept
(FPC) for floodplain-large river systems based on their experience in both temperate

406 Chapter 11 Freshwater Swamps and Riparian Ecosystems

and tropical regions of the world (Fig. 11.11). They dispute the RCC as a gener-
alizable theory because: (1) most of the theory was developed from experience on
low-order temperate streams, and (2) the concept is mostly restricted to habitats that
are permanent and lotic. In the FPC, the pulsing of the river discharge is the major
force controlling biota in river floodplains, and lateral exchange between the flood-
plain and river channel and nutrient cycling within the floodplain “have more direct
impact on biota than nutrient spiraling discussed in the RCC” (Junk et al., 1989). The
FPC thus considers the river-floodplain exchange to be of enormous importance in
determining the productivity of both the river and the adjacent riparian zone. Alter-
nating dry and wet cycles optimize productivity of the littoral zone and the adjacent
forest, decomposition of all that is produced, and fish spawning and feeding.

Recommended Readings

Messina, M. G., and W. H. Conner, eds. 1998. Southern Forested Wetlands. Boca
Raton, FT: Lewis Publishers.

References

Anderson, C. J., and W. J. Mitsch. 2008. The influence of flood connectivity on bot-
tomland forest productivity in central Ohio, USA. Ohio Journal of Science 108
(2): 2-8.

Anderson, R. C., and J. White. 1970. A cypress swamp outlier in southern Illinois.
Illinois State Acad. Sci. Trans. 63: 6-13.

Batzer, D. P, and S. A. Wissinger. 1996. Ecology of insect communities in nontidal
wetlands. Annual Review of Entomology Al-. 75-100.

Bayley, P. B. 1995. Understanding large river-floodplain ecosystems. BioScience
45:153-158.

Brinson, M. M. 1977. Decomposition and nutrient exchange of litter in an alluvial
swamp forest. Ecology 58: 601-609.

Brinson, M. M., B. L. Swift R. C. Plantico, and J. S. Barclay. 1981. Riparian Ecosys-
tems: Their Ecology and Status, U.S. Fish and Wildlife Service, Biol. Serv. Pro£[.,
FWS/OBS-81/17, Washington, DC, 151pp.

Brown, C. A. 1984. Morphology and biology of cypress trees. In K. C. Ewel and H. T.

Odum, eds.. Cypress Swamps . University Presses of Florida, Gainesville.

Brown, S. L. 1978. A Gomparison of Gypress Ecosystems in the Landscape of Florida.

Ph.D. dissertation. University of Florida, 569 pp.

Brown, S. L. 1981. A comparison of the structure, primary productivity, and transpi-
ration of cypress ecosystems in Florida. Ecological Monographs 51: 403-427.
Brown, S. L., and A. E. Lugo. 1982. A comparison of structural and functional charac-
teristics of saltwater and freshwater forested wetiands. In B. Gopal, R. E. Turner,
R. G. Wetzel, and D. F. Whigham, eds. Wetlands: Ecolojjy and Manajjement.
National Institute of Ecology and International Scientific Publications, Jaipur,
India, pp. 109-130.

Brown, S. L., and D. L. Peterson. 1983. Structural characteristics and biomass pro-
duction of two Illinois bottomland forests. American Midland Naturalist 110:
107-117.

References 407

Carter, M. R., L. A. Bums, T. R. Cavinder, K. R. Dugger, R L. Fore, D. B. Hicks, H. L.
Revells, and T. W. Schmidt. 1973 . Ecosystem Analysis of the Bijj Cypress Swamp and
Estuaries. U.S. Environmental Protection Agency 904/9-74, Region IV, Adanta.

Casey, W. P, and K. C. Ewel. 2006. Patterns of succession in forested depressional
wetlands in North Florida, USA. Wetlands 26: 147-160.

Clewell, A. F., and D. B. Ward. 1987. White cedar in Florida and along the northern
Gulf Coast. In A.D. Laderman, ed., Atlantic White Cedar Wetlands. Westview
Press, Boulder, CO, pp. 69-82.

Cochran, M. 2001. Effect of hydrology on bottomland hardwood forest productiv-
ity in central Ohio (USA). Master’s thesis. Natural Resources, The Ohio State
University, Columbus.

Conner, W. H., and J. W. Day, Jr. 1976. Productivity and composition of a bald
cypress-water tupelo site and a bottomland hardwood site in a Louisiana swamp.
American Journal of Botany 6A. 1354-1364.

Conner, W. H., J. G. GosseUnk, and R. T. Parrondo. 1981 . “Gomparison of the vege-
tation of three Louisiana swamp sites with different flooding regimes.” American
Journal of Botany 68: 32-331.

Gonner, W. H., and J. W. Day, Jr. 1982. The ecology of forested wetlands in the
southeastern United States. In B. Gopal, R. E. Turner, R. G. Wetzel, and D. F.
Whigham, eds.. Wetlands: Ecology and Manapfement . National Institute of Ecol-
ogy and International Scientific Publications, Jaipur, India, pp. 69-87.

Gonner, W. H., I. Mihalia, and J. Wolfe. 2002. Tree community structure and changes
from 1987 to 1999 in three Louisiana and three South Garolina forested wetlands.
Wetlands 22\ 58-70.

Dabel, G. V., and F. P. Day, Jr. 1977. Structural composition of four plant commu-
nities in the Great Dismal Swamp, Virginia. Torrey Botanical Club Bulletin 104:
352-360.

Day, F. P. 1984. Biomass and litter accumulation in the Great Dismal Swamp.
In K. G. Ewel and H. T. Odum, eds.. Cypress Swamps., University of Florida
Press, Gainesville, pp. 386-392.

Day, J. W., Jr., T. J. Butler, and W. G. Gonner. 1977. Productivity and nutrient export
studies in a cypress swamp and lake system in Louisiana. In M. Wiley, ed.. Estu-
arine Processes, vol. II, Academic Press, New York, pp. 255-269.

Day, J. W., Jr., J. Ko, J., Rybczyk, D. Sabins, R. Bean, G. Berthelot, G. Brantley,
L. Gardoch, W. Gonner, J. N. Day, A. J. Englande, S. Feagley, E. Hyfield, R. Lane,
J. Lindsey, J. Mistich, E. Reyes, and R. Twilley. 2004. The use of wedands in
the Mississippi delta for wastewater assimilation: A review. Ocean and Coastal
Manapiement 47\ 671-691.

DeGraaf, R. M. and D. D. Rudis. 1986. New England Wildlife: Habitat Natural His-
tory, and Disturbance. General Technical Report NE-108, Northeastern Forest
Experiment Station, Forest Service, U.S. Department of Agriculture, Washington,
DG, 491 pp.

DeGraaf, R. M., and D. D. Rudis. 1990. Herpetofaunal species composition and
relative abundance among three New England forest types. Forest Ecology and
Manapiement 32: 155-165.

408 Chapter 11 Freshwater Swamps and Riparian Ecosystems

Denny, G. C., and M. A. Arnold. 2007. Taxonomy and nomenclature of baldcypress,
pondcypress, and Montezuma cypress: One, two, or three species.^ HortTechnolojjy
(Jan-Mar 2007) 17(1): 125-127.

Dorge, C. L., W. J. Mitsch, and J. R. Wiemhoff. 1984. Cypress wedands in southern
Illinois. In K. C. Ewel and H. T. Odum, eds.. Cypress Swamps^ University Presses
of Florida, Gainesville, pp. 393-404.

Duever, M. J., J. E. Carlson, and L. A. Riopelle. 1984. Corkscrew Swamp: A virgin
cypress strand. In K. C. Ewel and H. T. Odum, eds.. Cypress Swamps. University
Presses of Florida, Gainesville, pp. 334-348.

Ewel, K. C., and W. J. Mitsch. 1978. The effects of fire on species composition in
cypress dome ecosystems. Florida Scientist 41: 25-31.

Ewel, K. C., and H. T. Odum, eds. 1984. Cypress Swamps. University Presses of
Florida, Gainesville, EL.

Ewel, K C., and L. P. Wickenheiser. 1988. Effects of swamp size on growth rates
of cypress {Taxodium distichum) trees. American Midland Naturalist 120:
362-370.

Golet, F. C., and D. J. Lowry. 1987. Water regimes and tree growth in Rhode Island
Atlantic white cedar swamps. In A. D. Laderman, ed., Atlantic White Cedar Wet-
lands. Westview Press, Boulder, CO, pp. 91-110.

Golet, F. C., A. J. K. Calhoun, W. R. DeRagon, D. J. Lowry, and A. J. Gold. 1993.
Ecology of Red Maple Swamps in the Glaciated Northeast: A Community Profile.
Biolojjical Report 12, U.S. Fish and Wildlife Service, Washington, DC. 151 pp.

Gomez, M. M., and F. P. Day, Jr. 1982. Litter nutrient content and production in the
Great Dismal Swamp. American Journal of Botany 69: 1314-1321.

Heimburg, K. 1984. Hydrology of north-central Florida cypress domes. In K. C. Ewel
and H. T. Odum, eds.. Cypress Swamps . University Presses of Florida, Gainesville,
pp. 72-82.

Huenneke, L. F., and R. R. Sharitz. 1986. Microsite abundance and distribution of
woody seedlings in a South Carolina cypress-tupelo swamp. American Midland
Naturalist 115: 328-335.

Johnson, F. L., and D. T. Bell. 1976. Plant biomass and net primary production along
a flood-frequency gradient in a streamside forest. Castanea 41: 156-165.

Johnson, B. L., W. B. Richardson, and T. J. Naimo. 1995. Past, present, and future
concepts in large river ecology. BioScience 45: 134-141.

Junk, W. J., P. B. Bayley, and R. E. Sparks. 1989. The flood pulse concept in
river-floodplain systems. In D. P. Dodge, ed., Proceedinpis of the International
Larpie River Symposium. Special issue of the Journal of Canadian Fisheries and
Aquatic Sciences \Q6\ 11-127.

Kitchens, W. M., Jr., J. M. Dean, L. H. Stevenson, and J. M. Cooper. 1975. The Santee
Swamp as a nutrient sink. In F. G. Howell, J. B. Gentry, and M. H. Smith, eds..
Mineral Cycling in Southeastern Ecosystems, ERDA Symposium Series 740513.
U.S. Government Printing Office, Washington, DC, pp. 349-366.

References 409

Kurz, H., and D. Demaree. 1934. Cypress buttresses in relation to water and air.
Ecology 15: 36-41.

Kurz, H., and K. A. Wagner. 1953. Factors in cypress dome development. Ecology 34:
17-164.

Laderman, A. D. 1989. The ecology of the Adantic white cedar wedands: A commu-
nityprofile. Biology Reporters (7.21). U.S. Fish and Wildlife Service, Washington,
DC. 114 pp.

Laderman, A. D., ed. 1998. Coastally Restricted Eorests. Oxford University Press,
Oxford, UK, 334 pp.

Leopold, L. B., M. G. Wolman, and J. E. Miller. 1964. Fluvial Processes in Geomor-
phology. W. H. Freeman, San Francisco. 522 pp.

Litde, E. L., Jr. 1971. Atlas of United States Trees., vol. 1, Conifers and important hard-
woods, Misc. Pub. No. 1146, U.S. Department of Agriculture — Forest Service,
USGPO, Washington, DG.

Lockaby, B. G., and M. R. Walbridge. 1998. Biogeochemistry. In M. G. Messina and
W. H. Gonner, eds.. Southern Forested Wetlands: Ecology and Management. Lewis
Publishers, Boca Raton, FL, pp. 149-172.

Megonigal, J. P, and F. P. Day, Jr. 1988. Organic matter dynamics in four seasonally
flooded forest communities of the Dismal Swamp. American Journal of Botany
75: 1334-1343.

Megonigal, J. P, W. H. Gonner, S. Kroeger, and R. R. Sharitz. 1997. Aboveground
produedon in southeastern floodplain forests: A test of the subsidy-stress hypoth-
esis. Ecology 78: 370-384.

Minshall, G. W., R. G. Peterson, K. W. Gummins, T. L. Bott, J. R. Sedall, G. E.
Gushing, and R. L. Vannote. 1983. Inter-biome comparison of stream ecosystem
dynamics. Ecological Monographs 53-1-25.

Minshall, G. W., K. W. Gummins, R. G. Peterson, G. E. Gushing, D. A. Bruins, J. R.
Sedall, and R. L. Vannote. 1985. Development in stream ecosystem theory. Cana-
dian Journal of Fisheries and Aquatic Sciences 37:130-137.

Mitsch, W. J. 1979. Interacdons between a riparian swamp and a river in southern
Illinois. In R. R. Johnson and J. F. McGomlick, tech, coords.. Strategies for the
Protection and Management of Floodplain Wetlands and Other Riparian Ecosys-
tems, Proceedings of a Symposium, Galaway Gardens, U.S. Forest Service General
Technical Report WO-12, Washington, DG, pp. 63-72.

Mitsch, W. J., and K. G. Ewel. 1979. Gomparadve biomass and growth of cypress in
Florida wetlands. American Midland Naturalist 101: 417-426.

Mitsch, W. J., G. L. Dorge, and J. R. Wiemhoff 1979. Ecosystem dynamics:
A phosphorus budget of an alluvial cypress swamp in southern Illinois. Ecology
60: 1116-1124.

Mitsch, W. J., and W. G. Rust. 1984. Tree growth responses to flooding in a bottom-
land forest in northeastern Illinois. Forest Science 30: 499-510.

410 Chapter 11 Freshwater Swamps and Riparian Ecosystems

Mitsch, W. J., J. R. Taylor, and K. B. Benson. 1991. Estimating primary productivity
of forested wetland communities in different hydrologic landscapes. Landscape
Ecolop/y 5: 75-92.

Mitsch, W. J., and J. G. Gosselrnk. 2000. Wetlands, 3rd ed. John Wiley & Sons, New
York, 920 pp.

Mitsch, W. J., and J. W. Day, Jr. 2004. Thinking big with whole ecosystem studies
and ecosystem restoration — A legacy of H.T Odum. Ecolopiical ModelUnp; 178:
133-155.

Mitsch, W. J., and S. E. Jorgensen. 2004. Ecoloptical Enptineerinpt and Ecosystem
Restoration. John Wiley and Sons, Hoboken, NJ. 411pp.

Motzkin, G., W. A. Patterson, and N.E.R. Drake. 1993. Fire history and vegeta-
tion dynamics of a Chamaecyparis thyoides v/cth.nd on Gape God, Massachusetts.
Journal of Ecology 81: 391-402.

Mulholland, P. J. 1979. Organic Garbon in a Swamp-Stream Ecosystem and Export
by Streams in Eastern North Garolina. Ph.D. dissertation. University of North
Garolina, Ghapel Hill.

The Nature Gonservancy. 1992. The forested wetlands of the Mississippi River: An
ecosystem in crisis. The Nature Gonservancy, Baton Rouge, LA, 25 pp.

Nessel, J. K. 1978. Distribution and Dynamics of Organic Matter and Phospho-
rus in a Sewage Enriched Gypress Strand. Masters thesis. University of Florida,
Gainesville, EL, 159 pp.

Noe, G. P, and G. R. Hupp. 2005. Garbon, nitrogen, and phosphorus accumulation
in floodplains of Atlantic Goastal Plain rivers, USA. Ecological Applications 15:
1178-1190.

Odum, E. P, J. T. Finn, and E. H. Franz. 1979. Perturbation theory and the subsidy-
stress gradient. BioScience 29: 349-352.

Odum, H. T. 1982. Role of wetland ecosystems in the landscape of Florida in Ecosys-
tem Dynamics in Freshwater Wetlands and Shallow Water Bodies, vol. 11. D. O.
Logofet and N. K Luckyanov, eds., SGOPE and UNEP Workshop, Genter of
International Projects, Moscow, USSR, pp. 33-72.

Odum, H. T., K. G. Ewel, W. J. Mitsch, and J. W. Ordway. 1977. Recycling treated
sewage through cypress wetlands in Florida. In F. M. DTtri, ed.. Wastewater Ren-
ovation and Reuse. Marcel Dekker, Inc., New York, pp. 35-67.

Odum, W. E., E. P. Odum, and Odum, H. T. 1995. Nature’s pulsing paradigm. Estu-
aries 18: 547-555.

Penfound, W. T. 1952. Southern swamps and marshes. Botanical Review 18:
413-446.

Powell, S. W., and F. P. Day. 1991. Root production in four communities in the Great
Dismal Swamp. American Journal of Botany 78: 288-297.

Rivera-Monroy, V. H., B, Branoff, E, Meselhe, A. McGorquodale, M. Dortch, G. D.
Steyer, J. Visser, and H. Wang. 2013. Landscape -level estimation of nitrogen
removal in coastal Louisiana wetlands: Potential sinks under different restoration
scenarios. Journal of Coastal Research 67(spl): 75-87.

Robertson, A. 1., P. Y. Bacon, and G. Heagney. 2001. The response of floodplain
primary production to flood frequency and timing. Journal of Applied Ecology 38:
126-136.

References 411

Schlesinger, W. H. 1978. Community structure, dynamics, and nutrient ecology in
the Okefenokee Cypress Swamp-Forest. Ecological Monographs 43-65.

Schneider, R. L., and R. R. Sharitz. 1986. Seed bank dynamics in a southeastern
riverine swamp. American Journal of Botany 73: 1022-1030.

Stromberg, J. C. 2001. Influence of stream flow regime and temperature on growth
rate of the riparian tree, Platanus wripfhtii., in Arizona. Freshwater Biology 46:
227-239.

Thorp, J. H., E. M. McEwan, M. F. Flynn, and F. R. Hauer. 1985. Invertebrate col-
onization of submerged wood in a cypress-tupelo swamp and blackwater stream.
American Midland Naturalist 113: 56-68.

Vannote, R. L., G. W. Minshall, K W. Cummins, J. R. Sedell, and C. E. Cushing.
1980. The river continuum concept. Canadian Journal of Fisheries and Aquatic
Sciences 37:130-137.

Vernon, R. O. 1947. Cypress domes. Science 105: 97-99.

Villa, J. A., and W. J. Mitsch 2015. Carbon sequestration in different wedand plant
communities in Southwest Florida. International Journal for Biodiversity Science,
Ecosystems Services and Manapiement. doi.org/10. 1080/21 5 13732. 20 14. 973909

Watts, A. C. 2013. Organic soil combustion in cypress swamps: Moisture effects and
landscape implications for carbon release. Forest Ecology and Management 294C:
178-187.

Watts, A. C., L. N. Kobziar, and J. R. Snyder. 2012. Fire reinforces structure of pond-
cypress {Taxodium distichum var. imbricarium) domes in a wedand landscape.
Wetlands 32: 430M48.

Watts, A. C., and L. N. Kobziar. 2013. Smoldering combustion and ground fires:
Ecological effects and muld-scale significance. Fire Ecology 9: 124-132.

Wharton, C. H., W. M. Kitchens, E. C. Pendleton, and T. W. Sipe. 1982. The Ecology
of Bottomland Hardwood Swamps of the Southeast: A Community Profile. U.S. Fish
and Wildlife Service, Biological Services Program, FWS/OBS-81/37. 133 pp.

Whigham, D. F., and C. J. Richardson. 1988. Soil and plant chemistry of an Atlantic
white cedar wedand on the Inner Coastal Plain of Maryland. Canadian Journal
of Botany 66: 568-576.

Wright, R. B., B. G. Lockaby, and M. R. Walbridge. 2001. Phosphorus availability
in an ardficially flooded southeastern floodplain forest soil. Soil Science Society of
America Journal 65: 1293-1302.

Yarbro, L. A. 1983. The influence of hydrologic variations on phosphorus cycling and
retention in a swamp stream ecosystem. In T. D. Fontaine and S. M. Bartell, eds..
Dynamics ofFotic Ecosystems. Ann Arbor Science, Ann Arbor, Ml, pp. 223-245.

Northern peatland in Estonia

Chapter 12

Peatlands

Peatlands include bo^s and fens distributed primarily in the cool boreal zones of
the world where excess moisture is abundant. Bojjs and fens can be formed in
several ways, originating either from aquatic systems, as in flowthrough
succession or quakinp; bops, or from terrestrial systems, as with blanket bops.
Althouph many types of peatlands are identifiable, classification accordinp to
chemical conditions usually defines three types: (1) minerotrophic (true fens),

(2) ombrotrophic (raised bops), and (3) transition (poor fens). Features of
many peatlands include acidity caused by cation exchanpe with mosses,
oxidation of sulfur compounds, and orpanic acids, low nutrients and primary
productivity, slow decomposition, adaptive nutrient-cyclinp pathways, and peat
accumulation. Several enerpy and nutrient budpets have been developed for
peatlands, with the 1942 enerpy budpet by Lindeman one of the first in
ecolopical sciences. Peatlands collectively are the larpest terrestrial storape of
carbon on the planet and are seen as potential sources of carbon to the
atmosphere if they are disturbed hydrolopically or if climate shifts.

As defined here, peatlands include the deep peat deposits of the boreal regions of the
world. Bogs and fens, the two major types of peatlands, occur as thick peat deposits in
old lake basins or as blankets the landscape. Many of these lake basins were formed by
the last glaciation, and the peatlands are considered to be a late stage of a filling-in pro-
cess. Bops are add peat deposits with no significant inflow or outflow of surface water
or groundwater and support acidophilic (acid-loving) vegetation, particularly mosses.
Fens, in contrast, are open peatland systems that generally receive some drainage from
surrounding mineral soils and are often covered by grasses, sedges, or reeds. They are
in many respects transitional between marshes and bogs. Fens are important as a suc-
cessional stage in the development of bogs and will be considered in that context here.

413

414 Chapter 12 Peatlands

Bogs and fens have been studied and described on a worldwide basis more exten-
sively than any other type of freshwater wetland; European and North American ecol-
ogy literatures are particularly rich in peatland studies. Peariands have been studied
because of their vast area in temperate climates, their unique biota and successional pat-
terns, their economic importance of peat as a fuel and soil conditioner, and, recenriy,
their importance in the global atmospheric carbon balance. Bogs have intrigued and
mystified many cultures for centuries because of such discoveries as the Iron Age “bog
people” of Scandinavia, who were preserved intact for up to 2,000 years in the non-
decomposing peat (see, e.g.. Glob, 1969; Coles and Coles, 1989).

Because bogs and other peatlands are ubiquitous in northern Europe and North
America, many definitions and words, some unfortunate, that now describe wedands
in general originated from bog terminology; there is also considerable confusion in
the use of terms such as fen, yivamp, moor, muskepj, heath, mire, marsh, hipfhmoor,
lowmoor, and peatland to describe these ecosystems. The words peatlands in general
and bop/s and fens 'm particular will be used in this chapter to include deep peat deposits,
mostly of the cold, northern, forested regions of North America and Eurasia. Peat
deposits also occur in warm temperate, subtropical, or tropical regions, and we refer
briefly to a major example of these, specifically the pocosins of the southeastern Coastal
Plain of United States.

Geographic Extent

Bogs and fens are distributed in cold temperate climates of high humidity, mosdy in the
Northern Hemisphere (Fig. 12.1), where precipitation exceeds evapotranspiration,
leading to moisture accumulation. There are also some peatlands in the Southern
Hemisphere in southern South America and in New Zealand. But the most extensive
areas of bogs and fens occur in Scandinavia, eastern Europe, western Siberia, Alaska,
and Canada. Major areas where a very large percentage of the landscape is peatland
include the Hudson Bay lowlands in Canada, the Fennoscandian Shield in northern
Europe, and the western Siberian lowland around the Ob and Irtysh rivers.

There are about 3.5 million km^ of peatlands in the world (Gorham, 1991).
This total includes 1.6 million km^ in the former Soviet Union (Botch et ah, 1995),
900,000 km^ ofwhich are in the Western Siberian lowlands (Kremenetski et ah, 2003).
Fennoscania has another 220,000 km^. In North America, Canada has approximately
1.10 million km^ of peatlands. Combining Canada’s total with an estimated 0.55 mil-
lion km^ of peatlands in the United States (including Alaska) (see Table 10.1), the
1.65 million km^ of northern peariands in North America represents almost one-half
of the world’s peariands.

Some peatlands, not illustrated in Figure 12.1, are found in the Southern Hemi-
sphere in southern South America and in New Zealand, but the size of these peariands
collectively is small compared to those in the Northern Hemisphere. The New Zealand
Land Resource Inventory (Cromarty and Scott, 1996) lists 3,113 km^ of wetlands in
the entire country, many of which are peariands. Included in that estimate are 439 km^

Geographic Extent 415

Figure 12.1 Area of abundant peatlands in the boreal zone (taiga) ef the Northern Hemi-
sphere. Peatlands are associated with boreai regions and their subaipine equivalents in
mountainous regions. South of the tree line (soiid line), woodiand tundra or subaipine areas
extend to the northern broken line. (After Wieder et al., 2006)

oipakihi (shallow-peat heathland) and another 356 km^ of forest-pakiha associations,
some of which support Sphagnum moss (Buxton et ah, 1996). Raised bogs in New
Zealand are not characterized by Sphapinum moss or ericaceous species common in
the Northern Hemisphere but by rushlike plants co-dominated by restiad (coming
from the family Restionaceae) bog species.

In the United States, peatlands dip into the conterminous United States from
northern Minnesota to northern Maine. In the northeast United States, bogs and fens
are common in Maine, New York, and Vermont. Peatlands are also fairly common in
the unglaciated Appalachian Mountains in West Virginia, such as in the Dolly Sods
Wilderness, which has plant species otherwise found only in sea-level eastern Canada.
Bogs and fens are found as far south as Illinois, Indiana, and Ohio in basins scoured
out by the Pleistocene glaciers in the north-central United States. The Middle Atlantic
Coastal Plain supports an expansive area of poorly drained peatlands, called pocosins.

416 Chapter 12 Peatlands

that are similar to more northern peatlands in that they are nutrient poor and domi-
nated by evergreen woody plants, such as bilberry, whortleberry, cranberry, heather,
and Labrador tea belonging to the Ericaceae or heath family. Pocosins once covered
12,000 km^, 70 percent of which were in North Carolina. Thirty-three percent of the
pocosins in North Carolina have been destroyed (Richardson, 2003).

Hydrology and Peatland Development

Two primary processes necessary for peatland development are a positive water bal-
ance and peat accumulation. First, a positive water balance, meaning that precipitation
is greater than evapotranspiration, is essential for peatland development and survival.
Water budgets for a fen, bog, and pocosin (see Chapter 4: “Wetland Hydrology”)
show that evapotranspiration is generally only 50 to 70 percent of precipitation. The
seasonal distribution of precipitation and excess water is important because peadands
require a humid environment year-round. In seasonally wet climates with cold winters,
such as in the midwestern United States south of Minnesota, Wisconsin, and Michi-
gan, peadands are not common where hot, dry summers persist. The southern limit
to bog species and, hence, to bogs is thought to be determined by the intensity of
solar radiadon in the summer months when precipitadon and humidity are otherwise
adequate to support bogs.

Some peadands also depend on local river systems to maintain their moisture
regime. Banaszuk and Kamocki (2008) reported on impaired fluvial peadands of the
Narew River valley in northeast Poland that have been impacted by a lower water
table, expansion of Phm^mites^ and soil subsidence. Reduced river discharges, linked
to a milder and drier climate over the last few decades, may be the cause.

A second requirement for peadand development is a surplus of peat producdon
over decomposidon, or accumuladon greater than decomposidon {A> D). Although
primary producdon is generally low in northern peadands compared to other ecosys-
tems, decomposidon is even more depressed, so peat accumulates. This is a nec-
essary condidon for the development of ombrotrophic bogs (see descripdon later
in this secdon). The continued development of the ecosystem is direcdy related to
the amount of surplus water and peat. For example, in a cool, moist maritime cli-
mate, peatlands can develop over almost any substrate, even on hill slopes. In con-
trast, in warm climates where both evapotranspiradon and decomposidon are elevated,
ombrotrophic peadands seldom develop even when a precipitadon surplus occurs.
Once formed, a bog is remarkably resistant to condidons that alter the water balance
and peat accumuladon. The perched water table, the water-holding capacity of the
peat, and its low pH create a microclimate that is stable under fairly wide environ-
mental fluctuations.

Given the condidons of water surplus and peat accumuladon, peadands develop
through terrestrialization (the infilling of shallow lakes) or paludification (the blan-
kedng of terrestrial ecosystems by overgrowth of peadand vegetadon). Three major
bog formadon processes are commonly seen: (I) quaking bog succession, (2) paludi-
ficadon, and (3) flowthrough succession.

Hydrology and Peatland Development 417

Figure 12.2 Typical profile of a quaking bog.

Quaking Bog Succession

Quaking bog succession is the classical process of terrestrialization, as described in
most introductory botany or limnology courses. Bog development in some lake basins
involves the filling in of the basin from the surface, creating a quaking bo£i (or Schwing-
moorm German; Fig. 12.2). Plant cover, only partially rooted in the basin bottom or
floating like a raft, gradually develops from the edges toward the middle of the lake. A
mat of reeds, sedges, grasses, and other herbaceous plants develops along the leading
edge of a floating mat of peat that is soon consolidated and dominated by Sphag-
num and other bog flora. The mat has all of the characteristics of a raised bog except
hydrologic isolation. The older peat is often colonized by shrubs and then forest trees
such as pine, tamarack, and spruce, which form uniform concentric rings around the
advancing floating mat.

These peatlands develop only in small lakes that have little wave action; they
receive their name from the quaking of the entire surface that can be caused by walk-
ing on the floating mat. After peat accumulates above the water table, isolating the
Sphagnum- dosmnSitcd flora from their nutrient supply, the bog becomes increasingly
nutrient poor. The development of a perched water table also isolates the peatland
from groundwater and nutrient renewal. The result is a classic concentric, or excentric,
raised ombro trophic raised bog.

The hydrology of raised bogs has been investigated and found to be more compli-
cated than originally thought, particularly for bogs that are on the edge of the boreal
zone. Studies in the Lake Agassiz region of Minnesota showed that bogs and fens
are part of a regional hydrology, with fens receiving groundwater and raised bogs
generally recharging groundwater (Fig. 12.3a). Raised bogs are normally assumed to
be disconnected from groundwater and fed only by precipitation. In a normal wet
climate, this pattern of bog hydrology is true as a downward flow of excess precipita-
tion deflects upwardly moving groundwater from mineral soil well below the surface

418 Chapter 12 Peatlands

beach ridge raised

b. Wet climate

I I precipitation-derived recharge minerotrophic groound water through transition

►groundwater How path

Figure 12.3 (a) Regional linkages between groundwater and raised bogs in the Lake
Agassiz region of Minnesota (area is approximately 10 km long and 30 m thick). Detailed
patterns of subsurface hydrology in the raised bogs are illustrated for (b) wet climate and
(c) dry climate. During wet periods, precipitation-derived recharge maintains a head that
flushes mineral-rich groundwater from the peat. During droughts, the water mound drops and
mineral-rich groundwater can move upward into the raised-bog peat. (After Siegel et al.,
1995; Glaser et al., 1997a)

(Fig. 12.3b). This accelerates peat accumulation, which, in turn, maintains the peat
and, hence, hydrologic mound in the landscape. During droughts, which can be fre-
quent events in peatlands on the edge of the boreal region, groundwater can move
upward to within 1 to 2 m of the peat surface (Fig. 12.3c) and dramatically influence
the peatland chemistry.

Paludification

A second pattern of bog evolution occurs when blanket bogs exceed basin boundaries
and encroach on formerly dry land. This process of paludification can be brought

Classification of Peatlands 419

about by climatic change, geomorphological change, beaver dams, logging of forests,
or the natural advancement of a peatland. Often the lower layers of peat compress
and become impermeable, causing a perched water table near the surface of what was
formerly mineral soil. This causes wet and acid conditions that kill or stunt trees and
allow only ombrotrophic bog species to exist. In some situations, the progression from
forest to bog can take place in only a few generations of trees (Heilman, 1968).

Flowthrough Succession

Intermediate between terrestrialization and paludification is flowthrough succession
(also termed topogmous development)^ in which the development of peatland modifies
the pattern of surface water flow. It involves the development of a bog from a lake basin
that originally had continuous inflow and outflow of surface water and groundwater.
As the peat continues to build, the major inflow of water may be diverted and areas
may develop that become inundated only during high rainfall. In the final stage, the
bog remains above the groundwater level and becomes a true ombrotrophic bog.

Classification of Peatlands

Peatlands develop within a complex interaction of climate, hydrology, topography,
chemistry, and vegetation development (succession). Because the physical and
biotic processes that form peatlands are complex and differ somewhat from region to
region, many different classification systems have been proposed over the past century
(Table 12. 1). Classification schemes have been based on at least seven features:

1. Floristics

2. Vegetation structure

3. Geomorphology (succession or development)

4. Hydrology

5. Chemistry

6. Stratigraphy

7. Peat characteristics

The last is used primarily for economic exploitation purposes. The other six are
closely interrelated, leading to classification schemes that combine several natural
features.

Landscape Classification

The developmental processes described above determine large-scale patterns of
peatland development that have been divided into the following four landscape
classifications.

I. Raised bojjs. These are peat deposits that fill entire bashis, are raised above
groundwater levels, and receive their major inputs of nutrients from precipitation.

420 Chapter 12 Peatlands

Table 12.1 Historical classification schemes for peatiands

Principal
Basis for
Ciassification

Minerai-

infiuenced

Peatiands

Transition

Peatiands

Precipitation-

Dominated

Peatlands

Reference

Topography

Fen

Bog or rasied bog

Generai use

Niedermoore (low
moor)

Ubergangsmoore

Hochmoore (high moor)

Weber (1907)

Hydroiogy

Geogenous

Limnogenous

Topogenous

Soiigenous

Ombrogenous

von Post and Graniund
(1926), Sjdrs (1948), Du
Rietz (1949), Damman
(1986)

Rheophiious

Soiigenous

Minerogenous

Transition

Ombrophilous

Ombrogenous

Kuiczynski (1949)

Waiter (1973)

Warner and Rubec (1997)

Water chemistry

Rich fen

Poor fen

Bog

Generai use; Sjors (1948)

Minerotrophic

Rheotrophic

Mesotrophic

Ombrotrophic

Moore and Beilamy (1974)
Moore and Bellamy (1974)

Nutrition

Nahrstoffreichere

Mitteireiche

Nahrstoffearme

Weber (1907)

Eutrophic

Mesotrophic

Oiigotrophic

Weber (1907), Pjavchenko
(1982)

Vegetation

Emergent or
forested fen

Transitionai

Moss-lichen or
forested bog

Cowardin et al. (1979),
Gorham and Janssens
(1992)

Source’. Revised from Bridgham et al. (1996)

These bogs are found primarily in the boreal and northern deciduous biomes. When
a concentric pattern of pools and peat communities forms around the most elevated
part of the bog, the bog is called a concentric domed bo£i. Bogs that form from previ-
ously separate basins on sloping land and form elongated hummocks and pools aligned
perpendicular to the slope are called excentric raised bo^s. In Europe, the former are
found near the Baltic Sea, and the latter are found primarily in the North Karelian
region of Finland.

2. Aapa peatlands. These wetiands, which also are called string bojjs and pat-
terned fens (Figs. 12.4), are found throughout the boreal region, often north of
the raised bog region. The dominant feature of these wetlands is the long, narrow
alignment of the higher peat hummocks {strings) that form ridges perpendicular
to the slope of the peadand and are separated by deep pools (Jlarks in Swedish).
In appearance, they resemble a hillside of terraced rice fields. The strings and flarks
develop perpendicular to the direction of the water flow. The pattern begins as a
series of scattered pools on the down slope, wetter edge of the water track. These
pools gradually coalesce into linear flarks. Peat accumulation in the adjacent strings
and the increasing impermeability of decomposing peat in the flarks accentuate
the pattern. Within the large water tracks, tree islands appear to be remnants of
continuous swamp forests that were replaced by sedge lawns in the expanding water
tracks.

(b)

Figure 12.4 Two oblique aerial images of string bogs in North America: (a) aerial pheto of
Cedarburg Bog in southwestern Wisconsin, showing a pattern ef parallel peat ridges (strings)
alternating with water-filled depressions (flarks) running diagonally across the lewer half of
the photegraph; (b) a string fen in Labrador, Canada. The strings stand nut because they
are vegetated with ericaceous shrubs and scrub trees over sphagnum moss, whereas flarks
are dominated by mosses and herbs or, in the case of the Canadian site, extensive stand-
ing water. (Photegraph (a) by G. Guntenspergen, reprinted with permission (b) by D. Wells,
reprinted by permission of C. Rubec and reprinted from Mitsch et al., 1994, p. 30, Fig. 30,
with permission from Elsevier Science)

421

422 Chapter 12 Peatlands

3. Paalsa bojjs. These bogs, found in the southern limit of the tundra biome,
are large plateaus of peat (20-100 m in breadth and length and 3 m high) generally
underlain by frozen peat and silt. The peat acts like an insulating blanket, actually
keeping the ground ice from thawing and allowing the southernmost appearance of
the discontinuous permafrost. In Canada, as much as 40 percent of the land area is
influenced by cyrogenic factors. When peat overlies frozen sediments, it influences
the pattern of the landscape. Many distinctive forms are similar to European aapa and
paalsa peatlands but are embedded in a continuous peat-covered landscape.

4. Blanket bo£S. These wetlands along the northwestern coast of Europe and
throughout the British Isles and are a result of paludification described above. The
favorable humid Atlantic climate allows the peat literally to “blanket” very large areas
far from the site of the original peat accumulation. Peat in these areas generally can
advance on slopes of up to 18 percent; extremes of 25 percent have been noted on
slopes covered by blanket bogs in western Ireland.

Chemistry-based Classification

The developmental processes described previously lead to increasing isolation of bogs
from surface and subsurface flows of both water and mineral nutrients. The degree
of hydrologic isolation of mires leads to a simple classification that is probably the
most frequently used today and is based on the degree to which the peatland receives
groundwater inflow as compared to only precipitation.

1 . Minerotrophic peatlands. These are true fens that receive water that has
passed through mineral soil. These peatlands generally have a high
groundwater level and occupy a low point of relief in a basin. They are also
referred to as rheotrophic peatlands rich fens m general use.

2. Mesotrophie peatlands. These peatlands are intermediate between mineral-
nourished (minerotrophic) and precipitation-dominated (ombrotrophic)
peadands. Another term used frequently for this class is transitional peatlands
or poor fens.

3. Ombrotrophic peatlands. These are the true raised bogs that have developed
peat layers higher than their surroundings and that receive nutrients and
other minerals exclusively by precipitation.

Another “trophic” classification of peatlands, found in older European literature
(Weber, 1907) and originally developed to classify peatlands and not lakes (Hutchin-
son, 1973), is the three-level trophic classification familiar to limnologists:

1. Eutrophic peatlands. Nutrient-rich peatlands; described by Weber (1907) as
Ndhrstoffreichere (eutrophe).

2. Mesotrophie peatlands. Same as before; described by Weber (1907) as
Mittelreiche (mesotrophe).

3 . Oligotrophic peatlands. Nutrient-poor peatlands; described by Weber (1907)
as Ndhrstoffearme (oligotrophe).

Classification of Peatlands 423

Hutchinson (1973) suggested that the process of peatland development could be
called ol{0otrophication. The terms mtrophic and oU^otrophic were applied to lakes and
their current limnological use by Naumann (1919) twelve years after Weber (1907)
applied the terms to peatlands Russian scientists such as Pjavchenko (1982) and Bazile-
vich and Tishkov (1982) continued to use this nomenclature for peatlands well into
the 1980s.

Bridgham et al. (1996) argued for caution in the use of the “-trophic” suffix
for classifying peatlands because the classic peatland gradient from minerotrophic to
ombrotrophic, characterized by surface water chemistry such as pH, conductivity,
and alkalinity, does not necessarily correlate with the eutrophic to oligotrophic
gradient, which is defined in terms of nutrient (e.g., nitrogen, phosphorus, and
potassium) availability. Bridgham et al. (1998) found evidence to suggest that there
was higher phosphorus availability in bogs and higher nitrogen availability in fens.
In other words, a strict correlation between measures of dissolved minerals and
available nutrients has never been established. They suggested resurrecting the terms
eutrophic and oligotrophic, which are rarely used in peadand literature today, because
they clearly refer to nutrients and not to other minerals. Such a resurrection did
not occur.

Hydrology-based Classification

Terms such as soligenous and ombropfenous actually refer to the hydrological and
topographic origins of the peatlands, not to the mineral conditions of the inflowing
water. A true hydrologic classification of peatlands based on the following two
categories is illustrated in Figure 12.5:

1. Ombro^enous peatlands. Open only to precipitation

2. Geopfenous peatlands. Open to outside hydrologic flows other than
precipitation:

a. Limno^enous peatlands. Develop along slow-flowing streams or lakes

b. Topo^enous peatlands. Develop in topographic depressions with at least
some regional groundwater flow

c. Soligenous peatlands. Develop with regional interflow and surface runoff

Canadian Classification

One of the more complete classifications developed for wedands in general and peat-
lands in pardcular is the Canadian Wedand Classificadon System (Warner and Rubec,
1997). This classificadon uses the terms minerotrophic and ombrotrophic in its water
chemistry classificadon and mineropienous and ombropfenous in its hydrological classifi-
cadon. Their simple classification of peatlands is:

1. Bo£. Peatland receiving water exclusively from precipitadon and not
influenced by groundwater; sphagnum-dominated vegetadon

2. Fen. Peatland receiving water rich in dissolved minerals; vegetadon cover
composed dominandy of graminoid species and brown mosses

424 Chapter 12 Peatlands

1 . Ombrogenous

2. Geogenous

a Limnogenous

b. Topogenous

c. Soligenous

Figure 12.5 Classification of peatlands based on hydrology. Two major categories are
geogenous peatlands, which are open to surface and groundwater flow, and ombrogenous
peatlands, which only receive precipitation. (After Damman, 1986)

3. Swamp. Peatland dominated by trees, shrubs, and forbs; waters rich in
dissolved minerals

Biogeochemistry

Soil and water chemistry are among the most important factors in the development
and structure of the peatland ecosystems. Factors such as pH, mineral concentration.

Biogeochemistry 425

available nutrients, and cation exchange capacity influence the vegetation types and
their productivity. Conversely, the plant communities influence the chemical prop-
erties of the soil water. In few wetland types is this interdependence so apparent
as in northern peadands. The major features of peatland biogeochemistry are dis-
cussed here.

Acidity and Exchangeable Cations

The pH of peatlands generally decreases as the organic content increases with the
development from a minerotrophic fen to an ombrotrophic bog (Fig. 12.6). Fens are
dominated by minerals from surrounding soils whereas bogs rely on a sparse supply
of minerals from precipitation. Therefore, as a fen develops into a bog, the supply
of metallic cations (Ca^’*', Mg^"'’, Na‘*‘, K'*') drops sharply. At the same time, as the
organic content of the peat increases because of the slowing of the decomposition
rate, the capacity of the soil to adsorb and exchange cations increases. These changes
lead to the domination by hydrogen ions, and the pH falls sharply. Fens, in contrast,
can range from slightly acidic (poor fens) to strongly alkaline (rich fens) depending on
groundwater flow rate and chemistry (Bedford and Godwin, 2003). Gorham (1967)
found that bogs in the English Lake District had a pH range of 3.8 to 4.4 compared
to noncalcareous fens, which had a pH range of 4.8 to 6.0. The Russian scientist
Pjavchenko (1982) assigned a pH range of 2.6 to 3.3 to oligotrophic bogs and a

Organic content, %

Figure 12.6 Soil pH as a function of organic content of peat soil. (After Gorham, 1967)

426 Chapter 12 Peatlands

range of 4.1 to 4.8 to mesotrophic bogs; a pH greater than 4.8 defined a eutrophic
(minerotrophic) fen.

As litde as 10 percent of the water supply from groundwater may change the pH of
a bog from 3.6 to 6.8, that is, from an ombrotrophic bog to a minerotrophic-rich fen.
In an upper peat of a Minnesota raised bog, pH and conductivity, both indicators of
mineral groundwater, increased dramatically in a drought year compared to a wet year.

The causes of bog acidity are not entirely clear, but five causes usually are cited
for the low pH:

1 . Cation exchange by Sphagnum. Cation exchange may be the most important
mechanism for the generation of acidity in peatlands. There is a direct
relationship between pH and the exchangeable hydrogen in peat, presumably
the result of the metabolic activity of the plants. Sphagnum peats have a high
exchangeable hydrogen and, consequently, a lower pH than
sedge-dominated peats.

2. Oxidation of sulfur compounds to sulfuric acid. Organic sulfur reserves in peat
may be oxidized to acidic compounds.

3. Atmospheric deposition. Sulfur deposition is a significant source of acidity,
depending on the oxidation state of the sulfur and the location of the bog.
Acid sources in precipitation and dry deposition are usually small except close
to sources of atmospheric pollution.

4. Biological uptake of cations by plants. Ions in the peat water are concentrated
by evaporation and are differentially absorbed by the mosses. This affects
acidity, for example, by the uptake of cations that are exchanged with plant
hydrogen ions to maintain the charge balance.

5. Buildup of organic acids by decomposition. Gorham et al. (1984) presented
evidence supporting this source of bog acidity. Organic acids help buffer the
system against the alkalinity of metallic cations brought in by rainfall and
local runoff.

A detailed hydrogen budget constructed for a Minnesota bog complex implicated
nutrient uptake as a major source of acidity (Table 12.2). About 15 percent of this
hydrogen budget represents ion exchange on the cell walls of Sphagnum. Most of this
acidity is neutralized by the release of cations during decomposition. Most of the rest
of the acidity is generated by organic acid production from fulvic and other acids that
result from the incomplete oxidation of organic matter and that buffer the pH of bogs
throughout the world at a value of about 4. In addition to decomposition, the major
source of alkalinity to neutralize the acids, the weathering of iron and aluminum and
runoff are major processes.

Limiting Nutrients

Bogs are exceedingly defrcient in available plant nutrients; fens that contain groundwa-
ter and surface water sources generally have considerably more nutrients. The paucity

Vegetation 427

Table 12.2 Acidity baiance for a Minnesota bog complex

Sources

Acidity (meq m ^ yr

Wet and dry deposition

-0.20 ± 10.7

Upiand runoff

-44.3 ± 18.6

Nutrient uptake

827 ± 248

Organic acid production

263 ± 50

Total

1,044

Sinks

Denitrification

12.2

Decomposition

784

Weathering

Outflow

142 ± 50

Total

1,044

meq = milliequivalents
Source: Urban et al. (1985)

of nutrients in bogs leads to two significant results, which are discussed in more detail
later in this chapter: (1) The productivity of nutrient-poor bogs is lower than that of
nutrient-rich fens; and (2) the characteristic plants, animals, and microbes have many
special adaptations to the low-nutrient conditions. Many studies have attempted to
find the ultimate limiting factor for bog primary productivity; this may be a com-
plex and academic question because all available nutrients are in short supply, and the
growing season is short and cool. Although calcium and potassium have been shown
to be limiting, nitrogen and phosphorus are the major limiting chemicals in bog and
fen productivity. When these nutrients are added in significant amounts to peatlands,
major vegetation shifts occur; with management such as mowing, the limiting fac-
tor can change from nitrogen to phosphorus. Bog formation in its latter stages is
essentially limited to nutrients brought in by precipitation. The effects on peatlands
of increased atmospheric sources of nitrogen throughout the developed world due to
fossil fuel burning has yet to be assessed adequately.

Vegetation

Bogs can be simple sphagnum moss peatlands, sphagnum-sedge peatlands,
sphagnum-shrub peatlands, bog forests, or any number or combination of aci-
dophilic plants. Mosses, primarily those of the genus Spha£inum^ are the most
important peat-building plants in bogs throughout their geographical range. Mosses
grow in cushionlike, spongy mats; water content is high, with water sometimes held
higher than it normally would be held by capillary action. Sphapfnum grows shoots
actively only in the surface layers (at a rate of about 1-10 cm annually); the lower
layers die off and convert to peat.

In North American peatlands, Sphapinum often grows in association with cotton
grass [Eriophorum va^inatum), various sedges {Carex spp.), and certain ericaceous

428 Chapter 12 Peatlands

shrubs, such as heather {Callutm vulgaris)^ leatherleaf {Chamaedaphne calyculata),
cranberry and blueberry ( Vaccinium spp.), and Labrador tea {Ledum palustre) . Trees
such as pine {Finns sylvestris)^ crowberry {Empetrum spp.), spruce {Picea spp.), and
tamarack {Larix spp.) are often found in bogs as stunted individuals that may be
scarcely 1 m high yet several hundred years old. Fens in the United States tend to
be dominated by a diverse community of plants that are distinct from boreal peat-
lands and typically include bryophytes, sedges ( Carex2,.nA other genera of Cyperaceae),
dicotyledonous herbs, and grasses (Amon et al., 2002; Bedford and Godwin, 2003).

Vegetation Patterns in a Minnesota Peatiand

Heinselman (1970) described seven vegetation associations in the Lake Agas-
siz peatlands of northern Minnesota that are typical of many of those in North
America. These occur in an intricate mosaic across the landscape, reflecting
the topography, chemistry, and previous history of the site. The vegetation
zones correspond closely to the underlying peat and to the present nutrient
status of the site. The seven major zones are:

1. Rich swamp forest. These forested wetlands form narrow bands In
very wet sites around the perimeter of peatlands. The canopy Is
dominated by northern red cedar {Thuja occidentalis); there are also
some species of ash {Fraxinus spp.), tamarack, and spruce. A shrub
layer of alder, Alnus rugosa, is often present, as are hummocks of
Sphagnum moss.

2. Poor swamp forest: These swamps, occurring downslope of the rich
swamp forests, are nutrient-poor ecosystems and are the most
common peatiand type in the Lake Agassiz region. Tamarack is
usually the dominant canopy tree, with bog birch {Betula pumila) and
leatherleaf In the understory and Sphagnum forming 0.3- to
0.6-m-high hummocks.

3. Cedar string bog and fen complex. This is similar to zone 2 except
that trees’-edge fens alternate with cedar {Thuja occidentalis) on the
bog ridges (strings) and treeless sedge (mostly Carex) in hollows
(flarks) between the ridges.

4. Larch string bog and fen. In this type of string bog, similar to zones 2
and 3, tamarack (Larix) dominates the bog ridges.

5. Black spruce-feathermoss forest. This type is a mature black spruce
{Picea mariana) forest that also contains a carpet of feathermoss
{Pieurozium) and other mosses. The trees are tall, dense, and of
similar ages. This peatiand occurs near the margins of ombrotrophic
bogs and generally does not have standing water.

Vegetation 429

6. Sphagnum-black spruce-beatherleaf bog forest. This is a widespread
wetiand type in northern North America. Stunted biack spruce is the
oniy tree, and there is a heavy shrub iayer of ieatherieaf, iaurei
(Ka/m/a spp.), and Labrador tea growing in iarge “piiiows" of
Sphagnum moss between spruce patches. This association is found
in convex reiief and is isoiated from minerai-bearing water.

7. Sphagnum-/eather/eaf-Kaimia-spruce heath. A continuous bianket of
Sphagnum moss is the most conspicuous feature; a iow shrub iayer
and stunted trees (usuaiiy biack spruce) are present in 5 to 10
percent of the area. Zones 6 and 7 occur on a raised bog.

in the water chemistry ciassification presented eariier in the chapter,
zones 1 through 4 wouid be ciassified as minerotrophic, zone 5 as transitionai,
zone 6 as semiombrotrophic, and zone 7 as ombrotrophic.

Although Spha£inum species are the characteristic peat-forming ground cover of
bogs, as sedges are of poor fens, there is a considerable overlap of species along the
chemical gradient from mineral poor to mineral rich and from low pH to high pH. In
a direct gradient analysis of vascular plants found in both bogs and fens in northern
Minnesota, the sedges Carex ol{0osperma and Eriophorum spissum decrease in cover
abundance with mineral enrichment of the peat, whereas tamarack {Larix laricina)
increases in abundance. Black spruce and the ericaceous shrubs Labrador tea and
Ieatherieaf, however, show dual peaks, indicating that their distribution is not con-
trolled by mineral water chemistry but by another gradient, such as water level or
possibly nitrogen or phosphorus availability.

Nicholson et al. (1996) investigated climatic and ecological gradients and how
they affected bryophyte distribution in the Mackenzie River basin in northwestern
Canada. They found that the most important variables that explained bryophyte
species distributions were water chemistry (Mg^"*", Ca^"*", H"''), height above the water
table, precipitation, and annual temperature. As a result of examining these gradients,
seven peatland groups were clustered from the original 82 sites in the basin: (1) poor
fens, (2) peat plateaus with thermokarst pools, (3) low-boreal bogs, (4) bogs and
peat plateaus without thermokarst pools, (5) low-boreal dry poor fens, (6) wet
moderate-rich fens, and (7) wet extremely rich fens. Thermokarst pools are features
of a permafrost landscape where permafrost thawing and subsequent ice melting
creates an uneven topography of mounds, sinkholes, caverns, and lake basins.

Locky et al. (2005) investigated black spruce {Picea mariana) swamps, fens, and
bogs in the southern boreal region of Manitoba, Canada. They emphasized the dis-
tinction between black spruce swamps and other peatlands in the region pointing
out their tendency to occur on gradual slopes, adjacent to water bodies, to contain
larger trees with significant cover, and to occur on shallower peat than other peatlands.
Chemically, these swamps are similar to moderate-rich fens.

430 Chapter 12 Peatlands

a 70

pH classes

Figure 12.7 Distribution of two bryophyte famiiies (Sphagnaceae and Ambiysteglaceae)
versus surface water pH fer 440 peatland plots across North America. Piets were counted if
they had at least one species of the famiiy cevering more than 25 percent of the tetal area.
The bimedal pattern suggests a classificatien of peatlands based en moss vegetation. (After
Gorham and Janssens, 1992)

In another study that attempted to relate vegetation directly to water chemistry,
Gorham and Janssens (1992) investigated two families of mosses (Sphagnaceae and
Ambiysteglaceae) at 440 sites across northern North America (Fig. 12.7). They found
a clear bimodal split in their occurrence, with Sphagnaceae most often in low-pH
peadands (mode pH, 4.0-4.25) and Amblystegiaceae in high-pH peadands (mode
pH, 6.76-7.0).

Black Spruce Peatlands

One of the dominant forested wedands in the world is the black spruce peadand
of the taiga of Canada and Alaska. Black spruce {Picea mariana), often growing
in association with tamarack {Larix laricina), is the tree species most associated
with forested peatlands in the boreal regions of North America. These wetlands are
esdmated to encompass about half of the palustrine shrub-scrub wetlands in Alaska

Vegetation 431

and cover an estimated 14 million ha in the state. Black spruce is mostly associated
with ombrotrophic (bog) rather than minerotrophic (fen) communities. In bogs,
it is found in associations with leatherleaf {Chamaedaphne calyculata), Labrador
tea {Ledum spp.), laurel {Kalmia latifolm), blueberry {Vaccinium spp.), and bog
rosemary (Andromeda poUfolm) . Sphapinum spp., of course, is found as ground cover
in these bogs. In Alaska, common associations include P. mariana with Vaccinium
uliginonsum, Ledum pfroenlandicum^ and feathermoss {Fleur ozium schreberi) and
P. mariana with Sphapfnum spp. and Cladina spp. (Post, 1996). In regions where
permafrost is prevalent, black spruce wetlands often occur in paalsa hummocks.

Carolina Pocosins

In contrast to the more northern peatiands, the woody vegetation of pocosins found
mostly near the southeastern United States coastal region of North and South Carolina
is dominated by evergreen trees and shrubs. Two broad community classes have been
identified, and their presence was related to fire frequency, soil type, and hydroperiod.
A Pinus-Ericalean (pine and heath shrub) community develops on deep organic soils
with long hydroperiods and frequent fire. Three associations within this community
are (1) pond pine {Pinus serotina) canopy with titi {Cyrilla racemiflora) and zenobia
{Zenobia pulverulenta) shrubs, (2) pond pine and loblolly bay {Gordonia lasianthus)
canopy with fetterbush {Lyonia lucida)^ and (3) pond pine canopy with titi and fet-
terbush shrubs. A conifer-hardwood community type is found on shallow organic soils
with slightly shorter hydroperiods. Two associations in this group are (1) pond pine
canopy with titi, fetterbush, red maple {Acer rubrum)^ and black gum {Nyssa sylvat-
ica) shrubs; and (2) pond pme and pond cypress {Taxodium distictium var. nutans)
canopy with red maple, titi, fetterbush, and black gum shrubs.

Peatland Adaptations

The vegetation in bogs and peatiands both controls and is controlled by its physical
and chemical environment. Some of the conditions for which adaptations are necessary
in peatiands are discussed here.

Waterlogging

Many bog plants, in common with wetland vegetation in general, have anatomical
and morphological adaptation to waterlogged anaerobic environments. These include
(I) the development of large intercellular spaces (aerenchyma or lacunae) for oxygen
supply, (2) reduced oxygen consumption, and (3) oxygen leakage from the roots to
produce a locally aerobic root environment. Sphagnum^ conversely, is morphologically
adapted to maintain waterlogging. The compact growth habit, overlapping leaves, and
rolled branch leaves form a wick that draws up water and holds it by capillarity. These
adaptations enable Sphagnum to hold water up to 15 to 23 times its dry weight.

432 Chapter 12 Peatlands

Acidification of the Externai interstitiai Water

Spha£inum has the unique ability to acidify its environment, probably through the
production of organic acids, especially polygalacturonic acids located on the cell walls.
The galacturonic acid residues in the cell walls increase the cation exchange capac-
ity to double that of other bryophytes. The adaptive significance of this peculiarity of
Sphagnum is unclear. The acid environment retards bacterial action and hence decom-
position, enabling peat accumulation despite low primary production rates. It has been
suggested that the high cation exchange capacity also enables the plant to maintain
a higher and more stable pH and cation concentration in the living cells than in the
surrounding water.

Adaptations to Nutrient Deficiency

Many bog plants have adaptations to the low nutrient supply that enable them to con-
serve and accumulate nutrients. Adaptations seen in bog plants include evergreenness;
sclerophylly, or the thickening of the plant epidermis to minimize grazing; uptake of
amino acids; and high root biomass. Some bog plants, notably cotton grass {Erio-
phorum spp.), translocate nutrients back to perennating organs prior to litterfall in
the autumn. These nutrient reserves are available for the following year’s growth
and seedling establishment. The roots of other bog plants penetrate deep into peat
zones to bring nutrients to the surface. Bog litter has been demonstrated to release
potassium and phosphorus, often the most limiting nutrients, more rapidly than other
nutrients, an adaptation that keeps these nutrients in the upper layers of peat. Many eri-
caceous plants have adapted to low concentrations of nitrogen by effectively utilizing
ammonium nitrogen in place of limited nitrate nitrogen under low-pH conditions, by
efficiently using nitrogen and even by utilizing organic nitrogen sources). Some bog
plants also carry out symbiotic nitrogen fixation. The bog myrtle {Myrica, gale) and
the alder develop root nodules characteristic of nitrogen fixers and have been shown
to fix atmospheric nitrogen in bog environments.

Carnivorous Plants

Another well-known adaptation to nutrient deficiency in bogs is the ability of carniv-
orous plants to trap and digest insects. This special feature is seen in several unique
insectivorous bog plants, including the pitcher plant {Sarracenia purpurea^ Fig. 12.8)
and sundew {Drosera spp.). A nutrient limitation study developed for Sarracenia
in Minnesota showed that although nutrient and insect additions did not increase
biomass, there were respective nutrient increases in the leaves of the plant. It was esti-
mated that insect capture accounts for approximately 10 percent of the plant’s nitrogen
and phosphorus needs (Chapin and Pastor, 1995). Pitcher plants are obligate host
to more invertebrate species than any other bog plant (Rymal and Folkerts, 1982).
In the water-filled pitcher in the plant sketch shown in Figure 12.8, a mosquito, a
midge, two sarcophagid flies, and a mite are suggested as captives. An aphid and three
moths feed exclusively on the tissue. Other insects are associated with other parts of
the plant.

Consumers 433

Figure 12.8 The pitcher piant {Sarracenia purpurea) including invertebrates that asseciate
with the plant. (After Damman and French, 1987)

Overgrowth by Peat Mosses

Many flowering plants are faced with the additional problem of being overgrown by
peat mosses as the mosses grow in depth and in area covered. Adapting plants must
raise their shoot bases by elongating their rhizomes or by developing adventitious
roots. Trees such as pine, birch, and spruce are often severely stunted because of the
moss growth and poor substrate; they grow better on bogs where the vertical growth
of moss has stopped.

Consumers

Mammals

The populations of animals in bogs are generally low because of the low productiv-
ity and the unpalatability of bog vegetation. Animal density is closely related to the
structural diversity of the peariand vegetation. For example, forested peatiands tend
to support the greatest number of small-mammal species, especially close to upland
habitats. Large mammals tend to roam over larger landscapes and are thus not spe-
cific to individual peatland types. In northern Minnesota and New England, moose
{Alces alces) are frequently found in small peadands. White-tailed deer {Odocoileus
vir£iinmnus) browse heavily in white cedar bogs in winter. Black bear ( Ursus ameri-
canus) use peatiands for escape cover and for food. The woodland caribou {Kangifer

434 Chapter 12 Peatlands

tarandus) was the largest mammal that was largely restricted to peatlands, but it disap-
peared from Minnesota in 1936, probably as a result of hunting pressure. Several large
predatory mammals have been reported to use or inhabit peatlands in North Amer-
ica, including the gray wolf ( Canis lupus), red wolf ( Canis rufus), puma {Felis concolor
cougar), and grizzly bear ( Ursus arctos horriblis) (Bedford and Godwin, 2003). Smaller
mammals closely associated with peatlands are beaver [Castor canadensis), lynx [Lynx
canadensis), fishers [Martes pennant), and snowshoe hares [Lepus americanus). The
beaver is a fairly recent import into Minnesota peatlands. It moved in along drainage
ditches, seldom penetrating deep into large peadands, but it has had a significant effect
on peatland flooding in northern Minnesota (Naiman et ah, 1991). Wet forests are
becoming the only habitats where wide-ranging mammals such as the black bear,
otter, and mink are found (Sharitz and Gibbons, 1982; L. D. Harris, 1989). This
is not so much because peadands are obligate habitats but because the clearing of
upland forests has forced the remaining populadon into the remaining large tracts of
forested wedands.

Amphibians and Reptiies

Glaser (1987) reported only seven species of amphibians and four species of reptiles
in northern Minnesota peadands. Acid waters below pH 5 appear to be the major
limiting factor in their ability to colonize bogs. Fens may have a more diverse array
of faunal species including several that are rare. In their review of fens of the United
States, Bedford and Godwin (2003) listed the bog turde [Clemmys muhlenbergii) and
eastern massasauga [Sisturus catenatus) as federally listed (threatened, endangered, or
considered for lisdng) repdles that use fens with high frequency. Other rare or uncom-
mon species associated with fens include mole salamanders [Ambystomia talpoideum)
and four-toed salamanders [Hemidactylium scutatum) that frequent small mountain
fens in the Appalachians (Murdock, 1994).

Birds

Many bird species are seen in peatlands during different times of the year (Fig. 12.9).
For example, Warner and Wells (1980) reported 70 species during the breeding sea-
son. Most of these are also common on upland sites, but a few depend on peadands
for survival. These include the sandhill crane [Grus candensis), great gray owl [Strix
nebulosa), short-eared owl [Asio flammeus) , sora [Porzana Carolina), and sharp-tailed
sparrow [Ammospiza caudacuta). In New England, as one moves from the Gana-
dian border south, the species change, but the new species have analogous positions
along the gradient (Fig. 12.9). Pocosins in the Southeast United States that support
mature pond pine [Pinus serotina) can be inhabited by the endangered red-cockaded
woodpecker [Picoides borealis) (Richardson 2003).

- Olive-sided flycatcher —
-Yellow-bellied flycatcher ■

- Gray jay

Spruce grouse

Black-backed woodpecker —
Boreal chickadee

Black-throated green warbler
White-winged crossbill

-Palm warbler

- Common yellowthroat

-Rusty blackbird

- Lincoln's sparrow

-Song sparrow

Tree swallow

Palm warbler

Savannah sparrow -

Black duck

—Ring-necked duck

VEGETATION

Black spruce
Tamarack

Ruffed grouse

Downy woodpecker

Great crested flycatcher
Eastern wood pewee —
Blue jay -

Rhodora
Bog Laural
Labrador Tea

Leatherieaf

Sedges
Pitcher plant
Cranberry

Sphagnum

Black-capped chickadee -
Red-eyed vireo

American redstart -

Eastern kingbird ■

Gray catbird

Common grackle

Yellow warbler

Common yellowthroat —

Swamp sparrow

Song sparrow

■ Rough-winged swallow -

■ Red-winged blackbird —

Wood duck

Red oak
Red maple

Cinnamon fern

Highbush blueberry
Poison sumac
White azalea

Leatherieaf

Sedges^

Pitcher plant
Cranberry Sphagnum

Figure 12.9 Comparison of bird distribution, typical of a lake-border bog in the northern and
southern parts ef the nertheastern United States. (After Damman and French, 1987)

435

436 Chapter 12 Peatlands

Ecosystem Function

The dynamics of peatlands reflect the realities of the harsh physical environment and
the scarcity of mineral nutrients. These conditions result in three major features:

1 . Bogs are systems of low primary productivity; fens are generally more
productive; Spha£inum mosses often dominate bogs, and other vegetation is
stunted in growth.

2. Bogs and fens are peat producers whose rates of accumulation are controlled
by a combination of complex hydrologic, chemical, and topographic factors.
This peat contains a great store of nutrients, most of it below the rooting
zone and thus unavailable to plants.

3. Low-nutrient peatlands in cold climates have developed several unique
pathways to obtain, conserve, and recycle nutrients. The amount of nutrients
in living biomass is small. Cycling is slow because of the low temperatures,
the nutrient deficiency of the litter, and the waterlogging of the substrate. It
is more active when peat production stagnates and when bogs receive
increased nutrient inputs.

Primary Productivity

Major organic inputs to bog systems come from the primary production of the vascular
plants, liverworts, mosses, and lichens. Among vascular plants, ericaceous shrubs and
sedges are the most important primary producers, and much of this production is
below ground. Mosses, especially sphagnum mosses, account for one-third to one-half
of the total production. Bogs and fens are usually less productive than most other
wetland types and are generally less productive than the climatic terrestrial ecosystems
in their region, about half that of a coniferous forest and a little more than a third
that of a deciduous forest (Table 12.3). According to Pjavchenko (1982), forested
peatlands produce a range of 260 to 400 g organic matter irT^ yr*\ with the low
value that of an om bro trophic bog and the high value that of a minero trophic fen.
Maimer (1975) cited a typical range of 400 to 500 g yr“^ for nonforested, raised
(ombro trophic) bogs in western Europe. In contrast, Lieth (1975) estimated the net
primary productivity in the boreal forest to average 500 g rcT^ yr“^ and in the tem-
perate forest to average 1,000 g yr^h The estimate for boreal forests probably
includes bog forests as well as upland forests. Annual above-ground productivity for
control plots in a northern Minnesota fen ranged from 87 ± 2 to 459 ± 34g
yr*^ over a four-year period with below-ground biomass (estimated at the last year) at
470 ± 79 g m-2 (Weltzin et al., 2005).

The measurement of the growth or primary productivity of Sphajjnum mosses
presents special problems not encountered in productivity measurements of other
plants. The upper stems of the plant elongate, and the lower portions gradually die
off, become litter, and eventually form peat. It is difficult to measure the sloughing
off of dead material to litter. It is equally hard to measure the biomass of the plant
at any one time because it is difficult to separate the living and dead material of the

Ecosystem Function 437

Table 12.3 Net primary productivity of peatiands in Europe and North America

Location

Type of
Peatland

Living
Biomass
(g dry wt xvr'^)

Net Primary
Productivity
(g dry wt
m-2 yr^'’-)

Reference

Europe

Western Europe

general nonwooded raised bog

1,200

400-500

Maimer (1975)

Western Europe

forested raised bog

3,700

340

Moore and Bellamy (1974)

Russia

eutrophic forested bog

9,700-11,000

400

Pjavchenko (1982)

mesotrophic forested bog

4,500-8,900

350

oligotrophic forested bog

2,200-3,600

260

Russia

mesotrophic Pinus-Sphagnum

Bog

8,500

393

Bazilevich and Tishkov (1982)

England

blanket bog

659 ± 53®

Forrest and Smith (1975)

England

blanket bog

635

Heal et al. (1975)

Ireland

blanket bog

316

Doyle (1973)

North America

Michigan

rich fen

34l‘>

Richardson et al. (1976)

Minnesota

forested peatland

15,941

1,014

Reiners (1972)

fen forest

9,808

651"

Manitoba

peatland bog

1,943

Reader and Stewart (1972)

Alberta

bog

280"

Szumigalski & Bayley (1996a)

poor fen

310"

moderate-rich fen

360"

lacustrine sedge fen

214"

extreme-rich fen

245"

Alberta

bog

390"

Thormann & Bayley (1997)

floating sedge fen

356"

lacustrine sedge fen

277"

riverine sedge fen

409"

Quebec

poor fen

114"

Bartsch & Moore (1985)

rich fen

335"

transitional fen

176"

®Mean ± standard deviation for seven sites.
'’Above ground only.

peat. The following two methods for measuring Sphagnum growth give comparable
results: (1) the use of “innate” time markers, such as certain anatomical or morpho-
logical features of the moss; and (2) the direct measurement of changes in weight.
Growth rates for Sphapinum determined by these two techniques generally fall in the
range of 300 to 800 g m“^ yr“^ (Table 12.4). Although Damman (1979) and Wieder
and Lang (1983) suggested that annual production should increase with decreasing
latitude, only S. mapfellanicum shows such a trend in Table 12.4. Evidently, local and
regional factors are more important than latitude.

It is generally expected that peadands are nutrient limited. However, this can
vary by plant species and community. Chapin et al. (2004) experimentally loaded

438 Chapter 12 Peatlands

Table 12.4 Comparison of selected data on production of Sphagnum species in order of decreasing latitude

Species®

Growth

(mm/yr)

Production
(g m-2 yr-i)

Latitude (N)

Location

Mean Annual
Precipitation
(mm)

Mean

Annuai

Temperature

(°C)

Source

fus

1.4-3.2

68° 22'

N Sweden

600

2.9

Rosswaii and Fieai (1975)

fus

—

250

63° 09'

S Finiand

532

3.5

Siivoia and Hanski (1979)

fus

—

220-290

63° 09'

S Finiand

532

3.5

K. Toionen, in Rochefort
etai. (1990)

fus

7-16

195

60° 62'

S Finiand

632

4-4.8

Pakarinen (1978)

mag

9.5

59° 50'

S Norway

1,250

5.9

Pedersen (1975)

ang

14.7

500

—

fus

9.8

56°05'

S Sweden

800

7.9

Damman (1978)

mag

7.8

100

—

mag

10-18

50-100

55°09'

Engiand

1,270

9.3

S. B. Chapman (1965)

ang

28-34

110-240

54° 46'

Engiand

1,980

7.4

Ciymo and Reddaway
(1971)

mag

14-15

230

54° 46'

Engiand

1,980

7.4

Forrest and Smith (1975)

ang

—

240-330

—

ang

38-43

110-440

54° 46'

Engiand

1,980

7.4

Ciymo (1970)

fus

6-7

75-83

54° 43'

Quebec

791

4.9

Bartsch and Moore (1985)

ang

4-17

19-127

—

T. R. Moore (1989)

fus

—

270

54° 28'

Engiand

1,375

7.4

Beiiamy and Rieley (1967)

fus

424-801

54° 20'

N Germany

714

8.4

Overbeck and Flappach
(1957)

mag

35-51

252-794

—

ang

120-160

488-1,656

—

fus

—

49° 53'

S Manitoba

517

2.5

Reader and Stewart (1971)

fus

17-24

240

49° 52'

NE Ontario

858

0.8

Pakarinen and Gorham
(1983)

fus

7-31

69-303

49° 40'

NW Ontario

714

2.6

Rochefort et ai. (1990)

mag

11-34

52-240

—

ang

20-39

97-198

—

mag

540

39°07'

West Virginia

1,330

7.9

Wieder and Lang (1983)

“'fus = Sphagnum fuscum\ mag = S. magellanicum] ang = S. angustifolium.
Source: Rochefort et al. (1990).

nitrogen, phosphorus,, and calcium carbonate (to raise pH) into a bog and fen in
northern Minnesota and examined plant community and species productivity in
response. In the bog, calcium carbonate and low additions of N (2 g N yr^^)
both increased above-ground net primary productivity (ANPP) while higher loads
of N (6g N actually inhibited growth. Fen graminoid growth responded

to increased P additions. Within both wetlands, there were variable responses to
experimental conditions among plant types and species. The authors surmised that
although nutrient availability is low in peatlands, this does not necessarily mean that
peatlands are nutrient limited.

Ecosystem Function 439

Decomposition

The accumulation of peat in bogs is determined by the production of litter (from
primary production) and the destruction of organic matter (decomposition). As with
primary production, the rate of decomposition in peat bogs is generally low because of
( 1 ) waterlogged conditions, (2 ) low temperatures, and ( 3 ) acid conditions. In fact, the
accumulation of peat in peatlands is due more to slow decomposition processes than
to net community productivity. Besides leading to peat accumulation, slow decompo-
sition leads to slower nutrient recycling in an already nutrient-limited system.

The pattern of Sphagnum decomposition is highest near the surface, where aerobic
conditions exist. By 20 cm depth, the rate is about one-fifth of that at the surface. This
pattern is caused by anaerobic conditions. The bulk of the organic decomposition that
does occur in peat bogs is by microorganisms, although the total numbers of bacteria
in these wetland soils are much fewer than in aerated soils. As pH decreases, the fungal
component of the decomposer food web becomes more important relative to bacte-
rial populations. Verhoeven et al. (1994) used a cotton-strip decomposition method
and found substantially lower decay rates in ombrogenous bogs compared to other
peatlands and mineral-soil wetlands. Total phosphorus (positive correlation) and soil
organic matter (negative correlation) explained 75 percent of the decay rates. Thus,
low nutrients and high organic matter (which keeps the soils reduced) contribute sig-
nificantly to the low decay rates in the bogs. Szumigalski and Bayley (1996b) found
the following progression of rates of decay of litter from peatlands in central Alberta:
Carex > Betula > mosses. The highest decomposition rates were with plant material
with the highest nitrogen content. Using a standard litter material of Carex lasiocarpa,
litter losses were in the following order:

poor fen > wooded rich fen > bog > open rich fen > sedge fen

In the same region, Bayley and Mewhort (2004) compared peat-accumulating
marshes and moderate-rich fens. Although these wetlands can appear similar, the
decomposition rates for fens were notably slower than marshes, which were attributed
to the higher water levels in the marshes.

There has been considerable speculation about factors that give rise to patterned
peatlands. The pattern of strings and flarks or hummocks and hollows, for example,
appears to be related to differential rates of peat accumulation. Rochefort et al.
(1990) determined that differential accumulation in a poor-fen system in north-
western Ontario, Canada, was caused more by differences in peat decomposition
rates than by differences in primary production rates. They found that, even though
the production rates of Sphapfnum in hummocks were generally about equal to the
rates in hollows or even lower than the rates in minerotrophic hollows, hummock
species had slower decomposition rates than those of hollow species. As a result, peat
accumulated faster on hummocks than in hollows, and hummocks may be expanding
at the expense of hollows.

440 Chapter 12 Peatlands

Peat Accumulation

The vertical accumulation rate of peat in bogs and fens is generally thought to be
between 20 and 80 cm/1,000 yr in European bogs (Moore and Bellamy, 1974),
although Cameron (1970) gave a range of 100 to 200 cm/1,000 yr for North
American bogs and Nichols (1983) reported an accumulation rate for peat of 150
to 200 cm/1,000 yr in warm, highly productive sites. Maimer (1975) described
a vertical growth rate of 50 to 100 cm/1,000 yr as typical for western Europe.
Assuming an average density of peat of 50 mg/mL, this rate is equivalent to a
peat accumulation rate of 25 to 50 g yr“^. Hemond (1980) estimated a rapid
accumulation rate of 430 cm/1,000 yr, eqivalent to 180 g yr^\ for Thoreau’s
Bog, Massachusetts.

Comparing Bog Energy Flow Estimates on Two Continents

One of the earliest energy budgets for any ecosystem was determined in the
classic study by Lindeman (1942) of Cedar Bog Lake, a small bog in northern
Minnesota (Fig. 12.10a). Although this energy budget is crude, the main fea-
tures have stood the test of time. Very little of the incoming radiation (<0.1
percent) is captured in photosynthesis. The two largest flows of organic energy
are to respiration (26 percent) and to storage as peat (70 percent). Energy
flow in the simplified food web is primarily to herbivores (13 percent), and
about 3.5 percent goes to decomposers. As the following two more recent
budget measurements show, the peat storage term is exceedingly high, and
decomposition losses are probably underestimated.

Bazilevich and Tishkov (1982) and Alexandrov et al. (1994) presented
a detailed energy flow through a mesotrophic (transition) bog in the Euro-
pean region of Russia (Fig. 12.10b). The bog is a sphagnum-pine (Sphagnum
girgenoshnii-Pinus sylvestris) community containing shrubs such as bilberry
(Vaccinium myrtillus). The total energy stored in the bog was estimated to
be in excess of 137 kg dry organic matter/m^, with dead organic matter (to
a depth of 0.6 m of peat) accounting for 94 percent of the storage. Living
biomass was 8.5kg/m^, or about 6 percent of the organic storage. Gross
primary productivity was 987 g yr^, or about 4,400 kcal m"^ yr^ (assum-
ing Ig organic matter = 4.5 kcal), with about 60 percent consumed by plant
respiration. The distribution of the net primary production came from trees
(39 percent), algae (28 percent), shrubs (21 percent), mosses and lichens (9
percent), and grasses (3 percent). The net primary production was primarily
consumed by decomposers; much less was consumed In grazing food webs.
Net accumulation of peat was 100 g m"^ yr^ (or about 450 kcal yr^).
Losses other than biotic decomposition, which accounted for most of the loss
of organic matter, were chemical oxidation and surface and subsurface flows.

Ecosystem Function 441

Figure 12.10 Diagrams of the energy flow in peatlands: (a) Cedar Bog Lake,
Minnesota, and (b) a Russian transition peatland. Flow in kcal yi^^- Flows in

(a) were originally published in calories whereas flows in (b) were published in gram
dry weight and converted to energy as 4.5 kcal g~^. ((a) After Lindeman, 1942;

(b) after Bazilevich and Tishkov, 1982; Alexandrov et. al., 1994)

Comparison of these two energy budgets from Russia and the United
States, carried out several decades apart, illustrates several points. First,
Lindeman’s Cedar Bog was approximately one-fourth as productive as Bazile-
vich and Tishkov’s Russian peatland, a possibility, given that the Russian site

442 Chapter 12 Peatlands

was described as transitional between a bog and a fen. Second, the Ameri-
can bog accumulated more peat than did the Russian peatland, results that
could reflect either the sophistication of the measuring techniques for the
time or what happens to peatlands as they transition from fens to true bogs.
Lindeman shows a high percentage of the productivity stored permanently as
peat. Assuming 50g/L as the density of peat, Lindeman’s peat accumulation
results in a high rate 350 cm/1,000 yr, whereas the Russian study has a more
reasonable 200 cm/1,000 yr.

Nutrient Budgets

Nitrogen

Nitrogen budgets for two peatlands — Thoreau’s Bog in Massachusetts and a perched
raised-bog complex in Minnesota — make an interesting comparison (Fig. 12.11).
Although total nitrogen input is comparable in both systems, the Minnesota
perched-bog system catches some runoff from surrounding uplands, whereas nitro-
gen fixation is the largest source of biologically active nitrogen in Thoreau’s Bog.
Otherwise, the budgets are remarkably similar despite the differences in bog type and
location. Both accumulate nitrogen in peat and lose a significant portion through
runoff Denitrification is an uncertain term.

Peatlands have often been identified as diverse wetlands. Increased nutrient load-
ing appears to reduce that diversity. Drexler and Bedford (2002) traced various nutri-
ents in a small peadand adjacent to a farm field in central New York. They found that
the farm field was a nutrient source with phosphorus and potassium loading com-
ing primarily from overland flow while nitrogen loading was delivered primarily via
groundwater. Peat concentrations of phosphorus and potassium, along with ground-
water flux of nitrate-nitrogen and NH4 -N, were all negatively correlated with plant
diversity. Nutrient loading promoted the growth of taller, monotypic stands of vascular
plants ( Calama£frostis canadensis, Carex lacustris, Epilobium hirsutum, and Typha lat-
ifoUa) at the expense of plant diversity.

Peadands appear to have the capacity for denitrificadon (Hemond, 1983), but
the magnitude in vivo is uncertain. Some recent studies have invesdgated nitrogen
dynamics in peadands. Wray and Bayley (2007) esdmated annual denitrificadon rates
for boreal marshes and fens in Alberta at 1 1 and 24 g N m“^ y^^^ , respecdvely, with N2
being the dominant product.

Bridgham et al. (2001) examined nutrient availability along an ombrotrophic-
minerotrophic gradient for 16 peadands in northern Minnesota. They found that N
availability generally increased along the gradient. They detected seasonal patterns
associated with N availability with NO3-N being more available in the summer
and NH4-N more available in the winter. Highest P availability was found in
minerotrophic swamps and beaver meadows, with the least availability found in fens
and bogs.

Ecosystem Function 443

Urban

surface

runoff

Atmospheric

Surface

runoff

Atmospheric

N fixation deposition Denitrification

Surface

runoff

Figure 12.11 Nitrogen budgets for two northern ombrotrophic bogs: (a) perched raised-bog
compiex in northern Minnesota and (b) smaii fioating-mat sphagnum bog (Thoreau’s Bog)
in Massachusetts. Values are in g N m~^ yr~^- (Data from Hemond, 1983; Urban and
Eisenreich, 1988)

Carbon

Carbon budgets for peatlands have drawn a great deal of interest, given the importance
of these ecosystems in global carbon dynamics (see Chapter 17: “Wedands and Cli-
mate Change” for more discussion on this topic). High-latitude peadands are known
to store tremendous amounts of carbon. In the Western Siberian lowlands, Kremenet-
ski et al. (2003) esdmated that peadands have an average peat depth of 2. 6 m and store
greater than 53.8 million metric tons of carbon. It is generally accepted that boreal
peadands were once carbon sinks, but there is litde consensus that they are contem-
porary sinks. Carbon budgets have been developed for small peadands and for large

444 Chapter 12 Peatlands

peatland-dominated watersheds (Rivers et al., 1998). The latter, a 1,500-km^ water-
shed in the Lake Agassiz peadands in Minnesota, illustrated that the peat watershed
had a net carbon storage of 12.7 g C yr^'^ but that there was a tenuous balance
between the watershed being a source and a sink of carbon. Inflows of carbon are
groundwater, precipitation, and net community productivity, whereas outtflows are
groundwater and surface flow and outgassing of methane. It was estimated from a
companion study (Glaser et al., 1997b) that peat is accumulating at a rate of 1 mm/yr
(100 cm/1,000 yr). This budget illustrates the importance of accurate hydrologic
measurements as well as biological productivity measurements in determining accu-
rate nutrient budgets. Mitsch et al. (2013) report a net carbon retention of 29 ± 8 g
C yr“^ for eight boreal peatlands where both carbon sequestration and methane
emissions were measured, an average rate considerably lower than net carbon retention
rates for temperate or tropical wetlands.

Methane emissions from peatlands have also been studied closely because of their
potential influence as greenhouse gases and the expansive store of carbon in peatlands.
Annual precipitation and water table position also seems to be a primary controller of
methane flux in peatlands (Heikkinen et al., 2002; Huttunen et al., 2003; Smemo
and Yavitt, 2006). Methane emission from a Finnish minerotrophic peatlands ranged
between 8 and 330 mg nr^ d“^ , and there was a positive correlation between methane
emission and water table level (Huttunen et al., 2003). Rubier et al. (2005) found
a similar seasonal average range of CH4 emission (10-350 mg irT^ d^^) but empha-
sized there was considerable spatial variability. They too found a significant relationship
between water table position and mean CH4 flux (Fig. 12.12). They noted the log
linear relationship between CH4 flux and water table position suggesting that only a
small increase in water table depth was needed to increase CH4 emission substantially.
In their review of the forested peatland literature, Trettin et al. (2006) concluded that

-20 -15 -10 -5 0 5 10

Water level above peat surface, cm

Figure 12.12 Methane flux as a function of mean water level in Canadian peatlands. Water
table indicates depth below (negative) or above (positive) the peat surface. Error bars are
standard deviations. (After Bubier et al., 2005)

References 445

decreased water table levels will result in decreased CH4 emission and increased CO2
emission from the peat surface. They emphasized this does not mean that these peat-
lands will necessarily decrease in their soil carbonpool as the difference may be made
up by changes in plant succession and increased productivity.

Recommended Readings

Rochefort, L., M. Strack, M. Poulin, J. S. Price, and C. Lavoie. 2012. Northern Peat-
lands. In: D. R. Batzer and A. H. Baldwin, eds.. Wetland Habitats of North Amer-
ica: Ecology and Conservation Concerns, pp. 119-134. Los Angeles: University of
California Press.

Wieder, R. K., and D. H. Vitt, eds., 2006. Boreal Peatlands Ecosystems. Berlin:
Springer-Verlag.

References

Alexandrov, G. A., N. I. Bazilevich, D. A. Logofet, A. A. Tishkov, and T. E. Shytikova.
1994. Conceptual and mathematical modeling of mater cycling in Tajozhny Log
Bob ecosystem (Russia). In B. C. Patten, ed. Wetlands and Shallow Continental
Water Bodies, Vol. 2. SPB Academic Publishing, The Hague, the Netherlands,
pp. 45-93.

Amon, J. P, C. A. Thompson, Q. J. Carpenter, and J. Miner. 2002. Temperate zone
fens of the glaciated midwestern USA. Wetlands 22: 301-317.

Banaszuk, P, and A. Kamocki. 2008. Effects of climatic fluctuations and land-use
changes on the hydrology of temperate fluviogenous mire. Ecological Engineering
32: 133-146.

Bartsch, I., and T. R. Moore. 1985. A preliminary investigation of primary produc-
tion and decomposition in four peadands near Schefferville, Quebec. Canadian
Journal of Botany 63: 1241-1248.

Bayley, S. E., and R. L. Mewhort. 2004. Plant community structure and functional
differences between marshes and fens in the southern boreal region of Alberta,
Canada. Wetlands 24: 277-294.

Bazilevich, N. L, and A. A. Tishkov. 1982. Conceptual balance model of chemical
element cycles in a mesotrophic bog ecosystem. In Ecosystem Dynamics in Eresh-
water Wetlands and Shallow Water Bodies, Vol. 2. SCOPE and UNEP, Centre of
International Projects, Moscow, USSR, pp. 236-272.

Bedford, B. L., and K. S. Godwin. 2003. Fens of the United States: Distribution, char-
acteristics, and scientific connection versus legal isolation. Wetlands 23: 608-629.

Bellamy, D. J., and J. Rieley. 1967. Some ecological statistics of a “miniature bog.”
Oikos 18: 33M:0.

Botch, M. S., K. I. Kobak, T. S. Vinson, and T. P. Kolchugina. 1995. Carbon pools
and accumulation in peatlands of the former Soviet Union. Global Bio^eochemical
Cycles 9: 37M:6.

446 Chapter 12 Peatlands

Bridgham, S. D., J. Pastor, J. A. Janssens, C. Chapin, and T. J. Malterer. 1996. Multi-
ple limiting gradients in peatlands: A call for a new paradigm. Wetlands 16: 45-65 .

Bridgham, S. D., K. Updegraff, and J. Pastor. 1998. Carbon, nitrogen, and phospho-
rus mineralization in northern wetlands. Ecology 79: 1545-1561.

Bubier, J., T. Moore, K. Savage, and P. Crill. 2005. A comparison of methane flux in
a boreal landscape between a dry and a wet year. Global Bio^eochemical Cycles 19:
GB1023.

Buxton, R. P, P. N. Johnson, and P. R. Espie. 1996. Sphagnum Research Programme:
The Ecological Effects of Commercial Harvesting. Science for Conservation: 25.
New Zealand Department of Conservation, Wellington, 33 pp.

Cameron, C. C. 1970. Peat deposits of northeastern Pennsylvania U.S. Geological
Survey Bull. 1317-A, 90 pp.

Ghapin, G. T., and J. Pastor. 1995. Nutrient limitations in the northern pitcher plant
Sarracenia purpurea. Canadian Journal of Botany 7 A. 728-734.

Ghapin, G. T., S. D. Bridgham, and J. Pastor. 2004. pH and nutrient effects on
above-ground net primary production in a Minnesota, USA bog and fen.
Wetlands 24i\ 186-201.

Ghapman, S. B. 1965. The ecology of Goom Rigg Moss, Northumberland: Some
water relations of the bog system. Journal of Ecology 53: 371-384.

Glymo, R. S. 1970. The growth of Sphagnum'. Methods of measurement. Journal of
Ecology 58: 13-49.

Glymo, R. S., and E. J. F. Reddaway. 1971. Productivity of Sphapfnum (bog-moss)
and peat accumulation. Hydrobiolo^ia (Bucharest) 12: 181-192.

Goles, B., and J. Goles. 1989. People of the Wetlands, Bojjs, Bodies and Lake-E)wellers.
Thames & Hudson, New York. 215 pp.

Gowardin, L. M., V. Garter, F. G. Golet, and E. T. LaRoe. 1979. Classification of
Wetlands and Deepwater Habitats of the United States. FWS/OBS-79/31. U.S.
Fish and Wildlife Service, Washington, DG. 103 pp.

Gromarty, D., and D. Scott. 1996. A Directory of Wetlands in New Zealand. Depart-
ment of Gonservation, Wellington, New Zealand.

Damman, A. W. H. 1978. Distribution and movement of elements in ombrotrophic
peat bogs. Oikos 30: 480-495.

Damman, A. W. H. 1979. Geographic patterns in peadand development in eastern
North America. In Proceedinpis of the International Symposium of Classification of
Peat and Peatlands. Hyytiala, Finland, International Peat Society, pp. 42-57.

Damman, A. W. H. 1986. Hydrology, development, and biogeochemistry of
ombrogenous peat bogs with special reference to nutrient relocation in a western
Newfoundland bog. Canadian Journal of Botany 64: 384-394.

References 447

Damman, A. W. H., and T. W. French. 1987. The ecology of peat bopjs of the glaciated
northeastern United States: A community profile. U.S. Fish and Wildlife Service,
Washington, DC. 100 pp.

Doyle, G. J. 1973. Primary production estimates of native blanket bog and meadow
vegetation growing on reclaimed peat at Glenamoy, Ireland. In L. G. Bass and
F. E. Wielgolasld, eds.. Primary Production and Production Processes, Tundra
Biome. Tundra Biome Steering Gommittee, Edmonton, Ganada, pp. 141-151.

Drexler, J. Z., and B. L. Bedford. 2002. Pathways of nutrient loading and impacts on
plant diversity in a New York peadand. Wetlands 22: 263-281.

Du Rietz, G. E. 1949. Huvudenheter och huvudgranser i Svensk myrvegetation.
Svensk Botanisk Tidkrift 43: 274-309.

Forrest, G. 1., and R. A. H. Smith. 1975. The productivity of a range of blanket bog
types in the Northern Pennines. Journal of Ecology 63: 173-202.

Glaser, P. H. 1987. The Ecolopjy of Patterned Boreal Peatlands of Northern Minnesota:
A Community Profile. U.S. Fish and Wildlife Service Biological Report 85 (7.14),
Washington, DG, 98 pp.

Glaser, P. H., D. 1. Siegel, E. A. Romanowicz, and Y. P. Shen. 1997a. Regional linkages
between raised bogs and the climate, groundwater, and landscape of northwestern
Minnesota. Journal of Ecology 85: 3-16.

Glaser, P. H., P. G. Bennett, D. 1. Siegel, and E. A. Romanowicz. 1997b.
Palaeo-reversals in groundwater flow and peatland development at Lost River,
Minnesota, USA. Holocene 6: 413M:21.

Glob, P. V. 1969. The Bop[ People: Iron Ap[e Man Preserved, trans. by R. Bruce-Mitford.
Gornell University Press, Ithaca, NY. 200 pp.

Gorham, E. 1967. Some chemical aspects of wetland ecology. Technical Memoran-
dum 90, Gommittee on Geotechnical Research, National Research Gouncil of
Ganada, pp. 2-38.

Gorham, E. 1991. Northern peatlands: Role in the carbon cycle and probable
responses to climatic warming. Ecological Applications 1: 182-195.

Gorham, E., and J. A. Janssens. 1992. Goncepts of fen and bog reexamined in relation
to bryophyte cover and the acidity of surface waters. Acta Societatis Botanicorum
Poloniae 61: 7-20.

Gorham, E., S. J. Eisenreich, J. Ford, and M. V. Santelmann. 1984. The chemistry of
bog waters. In W. Stumm, ed. Chemical Processes in Lakes. John Wiley 8c Sons,
New York, pp. 339-363.

Harris, L. D. 1989. The faunal significance of fragmentation of southeastern bot-
tomland forests. In D. D. Hook and R. Lea, eds.. Proceedings of the Symposium
the Forested Wetlands of the Southern United States, U.S. Department of Agricul-
ture, Forest Service, Southeastern Forest Experiment Station, Asheville, NG, pp.
126-134.

Heal, O. W., H. E. Jones, and J. B. Whittaker. 1975. Moore House, U.K. In T. Ross-
wall and O. W. Heal, eds.. Structure and function of tundra ecosystems. Ecological
Bulletin 20. Swedish Natural Science Research Gouncil, Stockholm, pp. 295-320.

448 Chapter 12 Peatlands

Heikkinen J. E. P., V. Elsakov, and P. J. Martikainen. 2002. Carbon dioxide and
methane dynamics and annual carbon budget in tundra wetland in NE Europe,
Russia. Global Biqgeochemical Cycles 16. doi:10.1029/2002GB0001930

Heilman, P. E. 1968. Relationship of availability of phosphorus and cations to forest
succession and bog formation in interior Alaska. Ecolojyy 49: 331-336.

Heinselman, M. L. 1970. Landscape evolution and peadand types, and the Lake Agas-
siz Peadands Natural Area, Minnesota. Ecological Monographs 40: 235-261.

Hemond, H. F. 1980. Biogeochemistry of Thoreau’s Bog, Concord, Mass. Ecological
Monographs 50: 507-526.

Hemond, H. F. 1983. The nitrogen budget of Thoreau’s Bog. Ecology 64: 99-109.

Hutchinson, G. E. 1973. Eutrophicadon: The scientific background of a contempo-
rary pracdcal problem. American Scientist 61: 269-279.

Huttunen, J. T., H. Nykanen, J. Turunen, and P. J. Mardkainen. 2003. Methane emis-
sions from natural peadands in the northern boreal zone in Finland, Fennoscan-
dia. Atmospheric Environment 37: 147-151.

Kulczynski, S. 1949. Peat bogs of Polesie. Acad. Pol. Sci. Mem.., Ser. B, No. 15.
356 pp.

Kremenetski, K. V., A. A. Velichko, O. K. Borisova, G. M. MacDonald, L. G. Smith, K.
E. Frey, and L. A. Orlova. 2003. Peatlands of Western Siberian lowlands: Gurrent
knowledge on zonadon, carbon content and Late Quaternary history. Quaternary
Science Review 22: 703-723.

Lieth, H. 1975. Primary producdon of the major units of the world. In H. Lieth and
R. H. Whitaker, eds.. Primary Productivity of the Biosphere. Springer-Verlag, New
York, pp. 203-215.

Lindeman, R. L. 1942. The trophic -dynamic aspect of ecology. Ecology 23: 399-418.

Locky, D. L., S. E. Bayley, and D. H. Vitt. 2005. The vegetadonal ecology of black
spruce swamps, fens, and bogs in southern boreal Manitoba, Ganada. Wetlands
25: 564-582.

Maimer, N. 1975. Development of bog mires. In A. D. Hasler, ed. Coupling of Land
and Water Systems . Ecology Studies 10. Springer-Verlag, New York, pp. 85-92.

Mitsch, W. J., R. H. Mitsch, and R. E. Turner. 1994. Wedands of the Old and New
Worlds — Ecology and Management. In W. J. Mitsch, ed. Global Wetlands: Old
and New. Elsevier, Amsterdam, the Netherlands, pp. 3-56.

Mitsch, W. J., B. Bernal, A. M. Nahlik, U. Mander, L. Zhang, G. J. Anderson, S. E.
Jorgensen, and H. Brix. 2013. Wedands, carbon, and climate change. Landscape
Ecology 28: 583-597.

Moore, P. D., and D. J. Bellamy. 1974. Peatlands. Sprhiger-Verlag, New York. 221 pp.

Moore, T. R. 1989. Growth and net producdon of Sphagnum at five fen sites subarctic
eastern Ganada. Canadian Journal of Botany 67: 1203-1207.

Murdock, N. A. 1994. Rare and endangered plants and animals of southern
Appalachian wedands. Water, Air, and Soil Pollution 77: 385-405.

Naiman, R. J., T. Manning, and G. A. Johnston. 1991. Beaver populadon fiuctuadons
and tropospheric methane emissions in boreal wedands. Biogeochemistry 12: 1-15.

References 449

Naumann, E. 1919. Nagra sypunkte angaende planktons okilogi. Med. Sarskild han-
syn till fytoplankton. Svensk Botanisk Tidkrift 13: 129-158.

Nichols, D. S. 1983. Capacity of natural wetlands to remove nutrients from wastew-
ater. Journal of the Water Pollution Control Federation 55: 495-505.

Nicholson, B. J., L. D. Gignac, and S. E. Bayley. 1996. Peatland distribution along
a north-south transect in the Mackenzie River Basin in relation to climatic and
environmental gradients. Ve^etatio 126: 119-133.

Overbeck, F., and H. Happach. 1957. Uber das Wachstum und den Wasserhaushalt
einiger Hochmoor Sphagnum, Flora 144: 335-402.

Pakarinen, P. 1978. Production and nutrient ecology of three Sphagnum species in
southern Finnish raised bogs, Annales Botanici Fennici 15: 15-26.

Pakarinen, P, and E. Gorham. 1983. Mineral Element Gomposition of Spha£inum
fuscum poLts collected from Minnesota, Manitoba, and Ontario. In Proceedings of
the International Symposium on Peat Utilization. Bemidji State University Center
for Environmental Studies, Bemidji, MN, pp. 417M:29.

Pedersen, A. 1975 . Growth measurements of five Sphagnum species in South Norway.
Norwejjian Journal of Botany 22: 277-284.

Pjavchenko, N. J. 1982. Bog ecosystems and their importance in nature. In D. O.
Logofet and N. K. Luckyanov, eds.. Proceedings of International Workshop on
Ecosystems Dynamics in Wetlands and Shallow Water Bodies, Vol. 1, SCOPE and
UNEP Workshop. Center for International Projects, Moscow, USSR, pp. 7-21.

Post, R.A. 1996. Functional Profile of Black Spruce Wetlands in Alaska. Report EPA
910/R-96-006, U.S. Environmental Protection Agency, Seatde, WA, 170 pp.

Reader, R J., and J. M. Stewart. 1971. Net primary productivity of the bog vegetation
in southeastern Manitoba. Can. J. Bot. 49:1471-1477.

Reader, R. J., and J. M. Stewart. 1972. The relationship between net primary pro-
duction and accumulation for a peadand in southeastern Manitoba. Ecology 53:
1024-1037.

Reiners, W. A. 1972. Structure and energetics of three Minnesota forests. Ecolopfical
Monographs A2\ 71-94.

Richardson, C. J. 2003. Pocosins: Hydrologically isolated or integrated wetiands on
the landscape.^ Wetlands 2Z\ 563-576..

Richardson, C. J., W. A. Wentz, J. P. M. Chamie, J. A. Kadlec, and D. L. Tilton. 1976.
Plant growth, nutrient accumulation and decomposition in a central Michigan
peadand used for effluent treatment. In D. L. Tilton, R. H. Kadlec, and C. J.
Richardson, eds.. Freshwater Wetlands and Sewage Effluent Disposal. University
of Michigan, Ann Arbor, pp. 77-117.

Rivers, J. S., D. 1. Siegel, L. S. Chasar, J. P. Chanton, P. H. Glaser, N. T. Roulet, and
J. M. McKenzie. 1998. A stochasdc appraisal of the annual carbon budget of a
large circumboreal peatland. Rapid River Watershed, northern Minnesota. Global
Biogeochemical Cycles 12: 715-727.

450 Chapter 12 Peatlands

Rochefort, L., D. H. Vitt, and S. E. Bayley. 1990. Growth, production, and decom-
position dynamics of Sphagnum under natural and experimentally acidified con-
ditions. Ecology 71: 1986-2000.

Rosswall, T., and O. W. Heal. 1975. Structure and function of tundra ecosystems.
Ecological Bulletin (Sweden) 20: 265-294.

Rymal, D. E., and G. W. Folkerts. 1982. Insects associated with pitcher plants (Sar-
racenia, salraceniaceae), and their relationship to pitcher plans conservation: A
review. Journal of Alabama Academy of Sciences 53: 131-151.

Sharitz, R. R., and J. W. Gibbons, eds. 1982. The Ecology of Southeastern Shrub Bogs
(Pocosins) and Carolina Bays: A Community Profile, U.S. Fish Wildlife Service,
Division of Biological Services, Washington, DG, FWS/OBS-82/04.

Siegel, D. 1., A. S. Reeve, R H. Glaser, and E. A. Romanowicz. 1995. Glimate-driven
flushing of pore water in peatlands. Nature 374: 531-533.

Silvola, J., and 1. Hanski. 1979. Garbon accumulation in a raised bog. Oecologia
(Berlin) 37: 285-295.

Sjors, H. 1948. Myrvegetation i bergslagen. Acta Phytogeographica SuecicalT. 1-299.

Smemo, K. A., and J. B. Yavitt. 2006. A multi-year perspective on methane cycling in
a shallow peat fen in central New York State, USA. Wetlands 26: 20-29.

Szumigalski, A. R., and S. E. Bayley. 1996a. Net above-ground primary produc-
tion along a bog-rich fen gradient in central Alberta, Ganada. Wetlands 16:
467-476.

Szumigalski, A. R., and S. E. Bayley. 1996b. Decomposition along a bog to rich fen
gradient in central Alberta, Ganada. Canadian Journal of Botany 74: 573-581.

Thormann, M. N., and S. E. Bayley. 1997. Aboveground net primary productivity
along a bog-fen-marsh gradient in southern boreal Alberta, Ganada. Ecoscience 4:
374-384.

Trettin, G. G., R. Lafiio, K. Minkkinen, and J. Laine. 2006. Influence of climate
change factors on carbon dynamics in northern forested peatlands. Canadian
Journal of Soil Science 86: 269-280.

Urban, N. R., S. J. Eisenreich, and E. Gorham. 1985. Proton cycling in bogs:
geographic variation in northeastern North America. In T. G. Hutchinson
and K. Meema, eds.. Proceedings of the NATO Advanced Research Workshop on
the Effects of Acid Deposition on Eorest, Wetland, and Agricultural Ecosystems.
Springer-Verlag, New York, pp. 577-598.

Urban, N. R., and S. J. Eisenreich. 1988. Nitrogen cycling in a forested Minnesota
bog, Canadian Journal of Botany 66: 435-449.

Verhoeven, J. T. A., D. F. Whigham, M. van Kerlchoven, J. O’Neill, and E. Maltby.
1994. Gomparative study of nutrient-related processes in geographically sepa-
rated wedands: Toward a science base for functional assessment procedures. In
W. J. Mitsch, ed.. Global Wetlands: Old World and New. Elsevier, Amsterdam, pp.
91-106

References 451

von Post, L., and E. Granlund. 1926. Sodra Sveriges Torvtillgangar. Sveriges Geolo-
giska Undersokning Set. G Avhandlingar och uppsater, No. 355. Arsbok 19 (2):
1-127.

Warner, B. G., and G. D. A. Rubec, eds. 1997. The Canadian Wetland Classification
System. National Wetlands Working Group, Wetlands Research Gentre, University
of Waterloo, Waterloo, Ontario, Ganada.

Warner, D., and D. Wells. 1980. Bird population structure and seasonal habitat use as
indicators of environment quality ofpeatlands. Minnesota Department of Natural
Resources, St. Paul, MN, 84 pp.

Weber, G. A. 1907. Aufbau und Vegetation der Moore Norddutschlands. Beibl. Bot.
Jahrbuchern. 90: 19-34.

Weltzin, J. F., J. K. Keller, S. D. Bridgham, J. Pastor, P. B. Allen, and J. Ghen.
2005. Litter controls plant community composition in a northern fen. Oikos 110:
537-546.

Wieder, R. K., and G. E. Lang. 1983. Net primary production of the dominant
bryophytes in a Sphagnum-dominated wedand in West Virginia. Bryolojjist 86:
280-286.

Wieder, R. K., D. H. Vitt, and B. W. Benscoter. 2006. Peatlands and the boreal forest.
In R. K. Wieder and D. H. Vitt, eds.. Boreal Peatland Ecosystems. Springer Verlag,
Berlin, pp. 1-8.

Wray, H. E., and S. E. Bayley. 2007. Denitrification rates in marsh fringes and fens in
two boreal peatlands in Alberta, Ganada. Wetlands 27\ 1036-1045.

Part IV

Traditional Wetland Management

Chapter 13

Wetland Classification

Wetlands have been classified since the early 1 900s, beginning with the peatland
classifications of Europe and North America. The U.S. Fish and Wildlife
Service has developed two major wetland classifications as the bases for wetland
inventories. The early (1956) classification described 20 wetland types based on
flooding depth, dominant forms of vegetation, and salinity regimes.
Classification of Wetiands and Deepwater Habitats of the United States,
published in 1979, uses a hierarchical approach based on systems, subsystems,
classes, subclasses, dominance types, and special modifiers to define wetlands and
deepwater habitats precisely. Canadian and international wetland
classification systems provide alternative systems that recognize 49 and 32
different wetland types, respectively. More recently, classifications based on
wetland function have been developed, including a functionally based approach
called the hydrogeomorphic classification. Wetland inventories are carried out
at many different scales with several different imageries and with both aircraft
and satellite platforms.

To deal realistically with wetiands on a regional scale, wetland scientists and managers
have found it necessary both to categorize the different types of wetlands that exist
and to determine their extent and distribution. The first of these activities is called
wetland classification, and the second is called a wetland inventory. Some of the ear-
liest efforts were undertaken to find wetlands that could be drained for human use;
later classifications and inventories centered on the desire to compare different types
of wetlands in a given region, often for their value to waterfowl. The protection of
multiple ecological values of wetlands came later; now it is the most common reason
for wetland classification and inventory. Recognition of wetland “value” has led some

455

456 Chapter 13 Wetland Classification

to now seek wetland classifications based on priorities for protection, with highest pro-
tection afforded to those wetlands with the greatest value. As with other techniques,
classifications and inventories are valuable only when the user is familiar with their
scope and limitations.

Our textbook uses a simple five-chapter classification of wetland ecosystems
(Chapters 8-12), divided into two major groups: (1) coastal — tidal marshes (salt
marshes and tidal freshwater marshes) and mangrove swamps, and (2) inland —
freshwater marshes, freshwater swamps, and peatlands (Tables 8.1 and 10.1). Other
types of wetlands, such as inland saline marshes, may fall between the cracks in this
simple wetland classification, but these five categories cover most wetlands currently
found in the world.

Why Do We Classify Wetlands?

Several attempts have been made to classify wetlands into categories that
follow their structural and functional characteristics. These classifications
depend on a well-understood general definition of wetlands (see Chapter 2:
“Wetland Definitions”), although a classification contains definitions of
individual wetland types. A primary goal of wetland classifications, according
to Cowardin et al. (1979), “is to impose boundaries on natural ecosystems
for the purposes of Inventory, evaluation, and management.” These authors
identified four major objectives of a classification system:

1. To describe ecological units that have certain homogeneous natural
attributes;

2. To arrange these units in a unified framework for the characterization
and description of wetlands, that will aid decisions about resource
management;

3. To identify classification units for inventory and mapping; and

4. To provide uniformity in concepts and terminology.

The first objective deals with the important task of grouping ecosystems
that have similar characteristics in much the same way that taxonomists cate-
gorize species in taxonomic groupings. The wetland attributes that are fre-
quently used to group and compare wetlands include the geomorphic and
hydrologic regime, vegetation physiognomic type, and plant and/or animal
species.

The second objective, to aid wetland managers, can be met In several
ways when wetlands are classified. Classifications (which are definitions of
different types of wetlands) enable wetland managers to deal with wetland reg-
ulation and protection consistently from region to region and from one time to
the next. Classifications also enable wetland managers to pay selectively more

Wetland Classifications 457

attention to those types of wetlands that are most threatened or functionally
the most valuable to a given region.

The third and fourth objectives, to provide consistency in the formulation
and use of inventories, mapping, concepts, and terminology, are also impor-
tant in wetland management. The use of consistent terms to define particular
types of wetlands is needed in the field of wetland science (see Chapter 2:
“Wetland Definitions”). These terms should then be applied uniformly to wet-
land inventories and mapping so that different regions can be compared and
so that there will be a common understanding of wetland types among wetland
scientists, wetland managers, and wetland owners.

Wetland Classifications

Peatland Classifications

Many of the earliest wetland classifications were undertaken for the northern peatlands
of Europe and North America. An early peadand classification in the United States,
developed by Davis (1907), described Michigan bogs according to three criteria: (1)
the landform on which the bog was established, such as shallow lake basins or deltas
of streams; (2 ) the method by which the bog was developed, such as from the bot-
tom up or from the shores inward; and (3) the surface vegetation, such as tamarack
or mosses. Based on the work of Weber (1907), Potonie (1908), Kulczynski (1949),
and others in Europe, Moore and Bellamy (1974) described seven types of peatlands
based on flowthrough conditions. Three general categories, called rheophilous, tran-
sition, and ombrophilous, describe the degree to which peatlands are influenced by
outside drainage. The more modern terminology is minerotrophic, transition, and
ombrotrophic peatlands (see Chapter 12: “Peatlands”). Most peatlands are limited to
northern temperate climes and do not include all or even most types of wedands in
North America. These classifications, however, served as models for more inclusive
classifications. They are significant because they combined the chemical and physi-
cal conditions of the wetland with the vegetation description to present a balanced
approach to wedand classificadon.

Circular 39 Classification

In the early 1950s, the U.S. Fish and Wildlife Service recognized the need for a
nadonal wedands inventory to determine “the distribudon, extent, and quality of the
remaining wedands in reladon to their value as wildlife habitat” (Shaw and Fredine,
1956). A classificadon was developed for that inventory (Mardn et ah, 1953), and the
results of both the inventory and the classificadon scheme were published in U.S. Fish
and Wildlife Circular 39 (Shaw and Fredine, 1956). Twenty types of wedands were
described under four major categories (Table 13.1).

Table 13.1 Early “Circular 39” wetland classification by U.S. Fish and Wildiife Service

Type Number Wetland Type

Site Characteristics

Iniand Ftesh Areas

Seasonally flooded

Soil covered with water or waterlogged during variable

basins or flats

periods, but well drained during much of the growing
season; in upland depressions and bottomlands

Fresh meadows

Without standing water during growing season;
waterlogged to within a few centimeters of surface

Shallow fresh

Soil waterlogged during growing season; often covered

marshes

with 15 cm or more of water

Deep fresh marshes

Soil covered with 15 cm to 1 m of water

Open fresh water

Water less than 2 m deep

Shrub swamps

Soil waterlogged; often covered with 15cm or more of
water

Wooded swamps

Soil waterlogged; often covered with 30 cm of water; along
sluggish streams, flat uplands, shallow lake basins

Bogs

Soil waterlogged; spongy covering of mosses

Iniand Saiine Areas

Saline flats

Flooded after periods of heavy precipitation; waterlogged
within few centimeters of surface during the growing
season

Saline marshes

Soil waterlogged during growing season; often covered
with 0.7 to Im of water; shallow lake basins

Open saline water

Permanent areas of shallow saline water; depth variable

Coastal Ftesh Areas

Shallow fresh

Soil waterlogged during growing season; at high tide, as

marshes

much as 15 cm of water; on landward side, deep
marshes along tidal rivers, sounds, deltas

Deep fresh marshes

At high tide, covered with 15 cm to Im water; along tidal
rivers and bays

Open fresh water

Shallow portions of open water along fresh tidal rivers and
sounds

Coastal Saline Areas

Salt flats

Soil waterlogged during growing season; sites occasionally
to fairly regularly covered by high tide; landward sides or
islands within salt meadows and marshes

Salt meadows

Soil waterlogged during growing season; rarely covered
with tide water; landward side of salt marshes

Irregularly flooded

Covered by wind tides at irregular intervals during the

salt marshes

growing season; along shores of nearly enclosed bays,
sounds, etc.

Regularly flooded

Covered at average high tide with 15 cm or more of water;

salt marshes

along open ocean and along sounds

Sounds and bays

Portions of saltwater sounds and bays shallow enough to
be diked and filled; all water landward from average
low-tide line

Mangrove swamps

Soil covered at average high tide with 15 cm to 1 m of
water; along coast of southern Florida

Source: Shaw and Fredine (1956)

458

Wetland Classifications 459

Types 1 through 8 are freshwater wetlands that include bottomland hardwood
forests (type 1), infrequently flooded meadows (type 2), freshwater nontidal marshes
(types 3 and 4), open water less than 2 m deep (type 5), shrub-scrub swamps (type 6),
forested swamps (type 7), and bogs (type 8). Types 9 through 11 are inland wetlands
that have saline soils. They are defined according to the degree of flooding. Types 12
through 14 are wetlands that, although freshwater, are close enough to the coast to
be influenced by tides. Types 15 through 20 are coastal wetlands that are influenced
by both saltwater and tidal action. These include salt flats and meadows (types 15
and 16), true salt marshes (types 17 and 18), open bays (type 19), and mangrove
swamps (type 20).

This wedand classification was the most widely used in the United States until
1979, when the current National Wetlands Inventory classification was adopted. The
earlier system is still referred to today by some wetland managers and is regarded by
many as elegandy simple compared with its successor. It primarily used the physiog-
nomy (life-forms) of vegetadon and the depth of flooding to idendfy the wedand
type. Salinity was the only chemical parameter used, and although wedand soils were
addressed in the Circular 39 publicadon, they were not used to define wedand types.

Coastal Wetland Classification

H. T. Odum et al. (1974) described coastal ecosystems by their major forcing func-
dons (e.g., seasonal programming of sunlight and temperature) and stresses (e.g.,
ice) (Fig. 13.1). Coastal wedand types in this classificadon include salt marshes and

s ^

IS 2

c g

O) ^

o ^

I- n

o «

(/) ^
0) -o

E>§

<D -
c c
0) g

? ‘c5

.9 N

<0 2
0) o

g §.

^ CA

Arctic systems

Naturally

Disturbed

systems

Temperate systems /

stressed

systems

with

wide

latitude

^ ^Tropical systems

■g

Increasing stress energies

Figure 13.1 Coastal ecosystem classification system based on latitude (and, hence, solar
energy) and major stresses. (After H. T. Odum et ai., 1974)

460 Chapter 13 Wetland Classification

mangrove swamps. Salt marshes, found in the type C category of natural temperate
ecosystems with seasonal programming, have “light tidal regimes” and “winter cold”
as forcing function and stress, respectively. Mangrove swamps are classified as type B
(natural tropical ecosystems) because they have abundant light, show litde stress, and
reflect little seasonal programming. Three additional classes, type A (naturally stressed
systems of wide latitudinal range), type D (natural arctic ecosystems with ice stress),
and type E (emerging new systems associated with human activity), were included in
this classification. The last class, which includes new systems formed by pollution, such
as pesticides and oil spills, is an interesting concept that still could be applied to other
wetland classihcations.

The United States Classification of Wetlands and Deepwater Habitats

The U.S. Fish and Wildlife Service began an inventory of the nation’s wetlands in
1974. Because this inventory was designed to fulfill several scientific and management
objectives, a new classihcation scheme, broader than the Circular 39 classihcation, was
developed and finally published in 1979 as a Classification ofi Wetlands and Deepwater
Habitats of the United States (Cowardin et ah, 1979). Because wetlands were found
to be continuous with deepwater ecosystems, both categories were addressed in this
classification. It is thus a comprehensive classification of all continental aquatic and
semiaquatic ecosystems. As described in that publication in 1979:

This classification, to be used in a new inventory of wetlands and deepwater
habitats of the United States, is intended to describe ecological taxa, arrange them in
a system useftil to resource managers, furnish units for mapping, and provide
uniformity of concepts and terms. Wetlands are defined by plants (hydrophytes),
soils (hydric soils), and frequency of flooding. Ecologically related areas of deep
water, traditionally not considered wetlands, are included in the classification as
deepwater habitats.

This classification is based on a hierarchical approach analogous to taxonomic
classifications used to identify plant and animal species. The first three levels of the clas-
sification hierarchy are given in Figure 13.2. The broadest level is systems: “a complex
of wetlands and deepwater habitats that share the influence of similar hydrologic, geo-
morphologic, chemical, or biological factors.” Thus, systems, subsystems, and classes
are based primarily on geologic and, to some extent, hydrologic considerations. Broad
vegetation types are included primarily at the class level, and, even here, the vegetation
types are generic (i.e., perennial, emergent, forested, scrub-shrub, or moss-Hchen).
Systems shown in Figure 13.2 include the following five:

1 . Marine. Open ocean overlying the continental shelf and its associated
high-energy coastline.

2. Estuarine. Deepwater tidal habitats and adjacent tidal wetlands that are
usually semienclosed by land but have open, partially obstructed, or sporadic
access to the ocean and in which ocean water is at least occasionally diluted
by freshwater runoff from the land.

System

Subsystem

Class

rock bottom
unconsolidated bottom
aquatic bed
reef

aquatic bed
reef

rocky shore
unconsolidated shore

rock bottom
unconsolidated bottom
aquatic bed
reef

aquatic bed
reef

stream bed
rocky shore
unconsolidated shore
emergent wetland
scrub-shrub wetland
forested wetland

rock bottom
unconsolidated bottom
aquatic bed
rocky shore
unconsolidated shore
emergent wetland

rock bottom
unconsolidated bottom
aquatic bed
rocky shore
unconsolidated shore

streambed

rock bottom
unconsolidated bottom
aquatic bed

rock bottom
unconsolidated bottom
aquatic bed
rocky shore
unconsolidated shore
emergent wetland

rock bottom
unconsolidated bottom
aquatic bed
unconsolidated shore
moss-lichen wetland
emergent wetland
scrub-shrub wetland
forested wetland

Figure 13.2 Current U.S. Fish and Wildlife Service wetland and deepwater habitat classifi-
cation hierarchy showing 5 major systems, 10 subsystems, and numerous ciasses. (After
Cowardin et ai., 1979)

461

462 Chapter 13 Wetland Classification

3. Riverine. Wetlands and deepwater habitats contained within a channel with
two exceptions: (1) wetlands dominated by trees, shrubs, persistent
emergents, emergent mosses, or lichens; and (2) deepwater habitats with
water containing ocean-derived salts in excess of 0.5 ppt.

4. Lacustrine. Wetlands and deepwater habitats with all of the following
characteristics: (1) situated in a topographic depression or a dammed river
channel; (2) lacking trees, shrubs, persistent emergents, emergent mosses, or
lichens with greater than 30 percent areal coverage; and (3) total area in
excess of 8 ha. Similar wedand and deepwater habitats totaling less than 8 ha
are also included in the lacustrine system when an active wave-formed or
bedrock shoreline feature makes up all or part of the boundary or when the
depth in the deepest part of the basin exceeds 2 m at low water.

5. Palustrine. All nontidal wedands dominated by trees, shrubs, persistent
emergents, emergent mosses, or lichens, and all such wedands that occur in
ddal areas where salinity stemming from ocean-derived salts is below 0.5 ppt.
It also includes wedands lacking such vegetation but with all of the following
characteristics: (1) area less than 8 ha; (2) lack of active wave-formed or
bedrock shoreline features; (3) water depth in the deepest part of the basin of
less than 2 m at low water; and (4) salinity stemming from ocean-derived salts
of less than 0.5 ppt.

Subsystems., as shown in Figure 13.2, give further defrnidon to the systems. These
include the following eight:

1 . Subtidal. Substrate condnuously submerged

2. Intertidal. Substrate exposed and flooded by tides, including the splash zone

3. Tidal. For riverine systems, gradient low and water velocity fluctuates under
ddal influence

4. Lower perennial. Riverine systems with continuous flow, low gradient, and no
ddal influence

5. Lfpper perennial. Riverine systems with continuous flow, high gradient, and
no ddal influence

6. Intermittent. Riverine systems in which water does not flow for part of the
year

7. Limnetic. All deepwater habitats in lakes

8 . Littoral. Wetland habitats of a lacustrine system that extends from shore to a
depth of 2 m below low water or to the maximum extent of nonpersistent
emergent plants

The class of a pardcular wedand or deepwater habitat describes the general appear-
ance of the ecosystem in terms of either the dominant vegetadon life form or the

Wetland Classifications 463

Upland Palustrine Upland Palustrine

Upland

Palustrine

Upland

seepage zone

a temporarily flooded
b seasonally flooded

d intermittently exposed
e permanently flooded

high water
- average water
■ low water

c semipermanently flooded f saturated

Figure 13.3 Features and examples of wetland classes and hydrologic modifiers in the
palustrine system. (After Cowardin et al., 1979)

substrate type. When more than 30 percent cover by vegetation is present, a veg-
etation class is used (e.g., shrub-scrub wetland). When less than 30 percent of the
substrate is covered by vegetation, then a substrate class is used (e.g., unconsolidated
bottom). The typical demarcation of many of the classes of the palustrine system is
shown in Figure 13.3.

Most inland wetlands fall into the palustrine system, in the classes moss-lichen,
emergent, scrub-shrub, or forested wetland. Coastal wetlands are classified in the same
classes within the estuarine system and intertidal subsystem. Only nonpersistent emer-
gent wetlands are classified into other systems.

Further descriptions of the wetlands and deepwater habitats are possible through
the use of subclasses^ dominance types, and modifiers. Subclasses such as “persistent”
and “nonpersistent” give further definition to a class such as emergent vegetation.
Type refers to a particular dominant plant species (e.g., bald cypress, Taxodium dis-
tichum, for a needle-leaved deciduous forested wetland) or a dominant sedentary or
sessile animal species (e.g., eastern oyster, Crassostrea vir^inica, for a mollusk reef).
Modifiers (Table 13.2) are used after classes and subclasses to describe more precisely
the water regime, the salinity, the pH, and the soil. For many wetiands, the description
of the environmental modifiers adds a great deal of information about their physical
and chemical characteristics. Unfortunately, those parameters are difficult to measure
consistendy in large-scale surveys such as inventories.

Table 13.2 Modifiers used in current wetiand and deepwater habitat classification by U.S.
Fish and Wildlife Service

Water Regime Modifiers (Tidai)

Subtidal — substrate permanently flooded with tidal water
Irregularly exposed — land suface exposed by tides less often than daily
Regularly flooded — alternately floods and exposes land surfaces at least daily
Irregularly flooded — land surface flooded less often than daily

Water Regime Modifiers (Nontidal)

Permanently flooded — water covers land surface throughout year in all years
Intermittently exposed — surface water present throughout year except in years of extreme drought
Semipermanently flooded — surface water persists throughout growing season in most years; when
surface water is absent, water table is at or near surface
Seasonally flooded — surface water is present for extended periods, especially in early growing season
but is absent by the end of the season

Saturated — substrate is saturated for extended periods during growing season but surface water is
seldom present

Temporarily flooded — surface water is present for brief periods during growing season but water table
is otherwise well below the soil surface

Intermittently flooded — substrate is usually exposed but surface water is present for variable periods
with no seasonal periodicity

Salinity Modifiers

Riverine, Lacustrine,

Marine and Estuarine

and Palustrine

Salinity (ppt)

Hyperhaline

Hypersaline

>40

Euhaline

Eusaline

30-40

Mixohaline (brackish)

Mixosaline

0.5-30

Polyhaline

Polysaline

18.0-30

Mesohaline

Mesosaline

5.0-18

Oligohaline

Oligosaline

0.5-5

Fresh

<0.5

pH Modifiers

Acid

pH <5.5

Circumneutral

pH 5.5-7.4

Alkaline

pH >7.4

Soil Material Modifiers

Mineral 1. <20% organic carbon and never saturated with water for more

than a few days, or

2. Saturated or artificially drained and has

a. <18% organic carbon if 60% or more is clay

b. <12% organic carbon if no clay

c. a proportional content of organic carbon between 12 and
18% if clay content is between 0 and 60%

Organic Other than mineral as described above

Source: Cowardin et al. (1979).

464

Wetland Classifications 465

Canadian Wetlands Classification System

The Canadian Wetland Classification System (Warner and Rubec, 1997) is designed
to be practical as well as hierarchical. Its three major features include:

1 . Classes. Based on natural features of the wetlands rather than on
interpretation for various uses. They have direct application to large wetland
regions. Wetland classes are recognized on the basis of properties that reflect
the overall “genetic origin” of the wetland system and the nature of the
wetland environment. Division into classes allows ready identification in the
field and delineation on maps. Classes are also convenient groupings for data
storage, retrieval, and interpretation.

2. Forms. Subdivisions of wetland classes based on surface morphology, water
type, and morphology characteristics of the underlying mineral soil. Some
forms are further subdivided into subforms. Forms are easily recognized
features of the landscape and are the basic wetland-mapping unit.

3 . Types. Subdivisions of wetland forms and subforms based on physiognomic
characteristics of the vegetation communities. They are comparable to the
modifiers used in the U.S. Fish and Wildlife Service classification system.
Types are most useful for evaluation of wetland values and benefits,
management for wetland hydrology and wildlife habitat, and conservation
and protection of rare and endangered species.

Currently, the system recognizes five wetland classes (bog, fen, swamp, marsh,
shallow-water marsh), 49 wedand forms, and 75 subforms, although geomorpholog-
ical, hydrologic, and chemical characteristics do not appear in the classification.

International Wetland Classification System

Table 13.3 compares the international Ramsar Convention classification system with
the U.S. and Canadian classification systems. The Ramsar system has 32 classes,
divided into a marine/coastal group and an inland group. Because it attempts to be
global, it has categories that neither the U.S. nor the Canadian systems have — for
example, underground karst systems and oases. The U.S. system, because it is
hierarchical, has fewer classes at the system and subsystem level but uses modifiers
to identify specific wetland types. The Canadian system also has only five classes but
with forms and subforms reaches a total of over 70 different categories by which to
classify wetlands. The Ramsar Convention itself is described in more detail in Chapter
15: “Wetland Laws and Protection.”

Hydrogeomorphic Wetland Classification

Mark Brinson (1993) developed a wetland classification system modeled after the
hydrogeomorphic (HGM) classification systems for mangroves (see Chapter 9:
“Mangrove Swamps”) and cypress swamps (see Chapter 11: “Freshwater Swamps
and Riparian Ecosystems”) that came from the H. T. Odum program at University of

Table 13.3 The Ramsar Convention International Wetland classification system compared to the U.S. Fish
and Wildlife wetland and deepwater habitat classification and the Canadian wetland classification system

Ramsar Convention Code Name

U.S. Fish and Wildlife System®

Canadian System*’

Marine/Coastal Wetlands

Marine water <6 m
Marine subtidal aquatic beds

Marine subtidal

Marine subtidal aquatic bed

Shallow (<2 m) water marsh

Coral reefs

Marine subtidal reef

—

Rocky marine shores

Marine intertidal rock bottom

—

Sandy shore or dune

Marine intertidal unconsolidated

—

Estuarine waters

Estuarine subtidal

Estuarine marsh, water

Intertidal flats

Estuarine intertidal unconsolidated bottom

Estuarine water, tidal water

Intertidal marshes

Estuarine intertidal emergent wetland

Tidal marsh

Intertidal forested wetland

Estuarine intertidal forested wetland

Tidal swamp

Coastal saline lagoon

Estuarine subtidal unconsolidated, saline

Estuarine water

Coastal fresh lagoon

Estuarine subtidal unconsolidated, fresh

Estuarine water

Zk(a) Marine/coastal karst

Inland Wetlands

Estuarine subtidal rocky shore

Permanent inland deltas

Riverine perennial
Estuarine delta marsh
Shallow riparian delta water

Riparian delta marsh

Permanent rivers/streams

Riverine perennial swamp, marsh

Shallow riparian water

Intermittent rivers/ streams

Riverine intermittent

—

Permanent fresh lakes
Riparian water (oxbows)

Lacustrine littoral or limnetic

Shallow lacustrine water

Intermittent fresh lakes

Lacustrine or riparian littoral

—

Permanent saline lakes

Lacustrine littoral unconsolidated, saline

—

Intermittent saline lakes

Lacustrine littoral intermittent, saline

—

Permanent saline
marshes/pools

Palustrine emergent wetland or
unconsolidated bottom, saline

Estuarine marsh, inland salt
swamp

Intermittent saline
marshes/pools

Palustrine emergent wetland or
unconsolidated bottom, intermittent

Spring, slope, or basin marsh

Permanent fresh marsh/ pools
(<8 ha)

Palustrine emergent wetland or
unconsolidated bottom, fresh

Shallow basin water, lacustrine
marsh

Intermittent fresh
marsh/ pools, inorganic soils

Palustrine emergent wetland, intermittently
flooded

Shallow basin water

Nonforested peatlands

Palustrine emergent wetland, persistent

Bogs, fens

Alpine wetlands

Palustrine emergent wetland, persistent

—

Tundra wetlands

Palustrine emergent wetland, persistent

Bogs, fens, shallow basin water

Shrub-dominated wetlands

Palustrine scrub-shrub wetland

Riparian, flat, slope, discharge
or mineral-rise swamp

Fresh forested wetlands on
inorganic soils

Palustrine forested wetland

Riparian swamp

Forested peatlands

Palustrine forested or scrub-shrub wetland

Flat bog, fiat or raised peatland
swamp

Freshwater springs, oases

—

Hummock marsh

Geothermal wetlands

—

Zg(b)

Inland karst systems,
underground

®Ramsar class can be further approximated with additional modifiers.

‘’Class and form indicated only; subforms can approximate Ramsar more closely. The term shallow in the wetland type indicates
the shallow marsh class.

466

Wetland Classifications 467

Figure 13.4 Basis of the hydregeomorphic (HGM) classification system: (a) geomerphic
settings (basin, fringe, and riverine) arranged around three core factors of hydroperiod,
hydrologic energy, and nutrient leveis; (b) the relative contribution of a combination of three
water sources — precipitation, groundwater discharge, and surface Inflew — to determining the
type of wetlands. (After Brinson, 1993)

Florida in the 1970s. It was designed to be used for evaluation of wedand functions
and is currently being used as a means of assessing the physical, chemical, and biolog-
ical functions of wetlands. It is useful for comparing the level of functional integrity
of wetlands within a functional class or for evaluating the impact of proposed human
activities on wetlands and mitigation alternatives (Fig. 13.4; Table 13.4).

The classification is based primarily on hydrodynamic differences as they function
within four geomorphic settings. Thus, the three core components of the classification
system are geomorphology, water source, and hydrodynamics (Table 13.4). Geomor-
phic setting is the topographic location of a wetland in the surrounding landscape.
Four geomorphic settings are identified: depressional, riverine, fringe, and extensive
peatiands. The first three are clearly related to the hydrologic setting. Extensive peat-
lands are different because the dominant influence on hydrology is biogenic accretion.
Water sources are precipitation, surface or near-surface flow, and groundwater dis-
charge (into a wetland). The term hydrodynamics rcicTS to the direction and strength
of water movement within a wetland. The three core features are heavily interdepen-
dent, so it is difficult to describe any one without the other two. Taken as a group,
the three core features may be pooled in 36 combinations, but because of the interde-
pendence, not all combinations are found in nature. It becomes clear when examining
Table 13.4 how interrelated the three core features are. The water entering a wetland
is seldom from only one of the three sources — ^precipitation, groundwater discharge,
and surface inflow — although ombrotrophic peat wetlands typically are dominated by
precipitation; mineral fens and seep wetlands, by groundwater discharge; and riverine
and fringe wetlands, by surface flows.

The core component hydrodynamics is an expression of the fluvial energy that
drives the system. This ranges from low-energy water table fluctuations typical

468 Chapter 13 Wetland Classification

Table 13.4 Functional classification of wetiands by geomorphology, water source, and
hydrodynamics

Core Component Description

Example

Geomorphic

Setting

Depressional

Extensive

peatlands

Riverine

Fringe

Water Source

Precipitation

Groundwater
discharge
Surface inflow

Hydrodynamics

Vertical

fluctuation

Unidirectional

flow

Bidirectional flow

Topographic location of a wetland in the
surrounding landscape

Wetlands in depressions that typically receive
most moisture from precipitation, hence often
ombrotrophic; found in dry and moist climates
Peat substrate isolates wetland from mineral
substrate; peat dominates movement and
storage of water and chemicals
Linear strips in landscape; subject
predominantly to unidirectional surface flow
Estuarine and lacustrine wetlands with
bidirectional surface flow

Relative importance of three main sources of
water to a wetland

Wetlands dominated by precipitation as the
primary water source; water level may be
variable because of evapotranspiration
Primary water source from regional or perched
mineral groundwater sources
Water source dominated by surface inflow

Motion of water and its capacity to do work
Vertical fluctuation of the water table resulting
from evapotranspiration and replacement by
precipitation or groundwater discharge
Unidirectional surface or near-surface flow;

velocity corresponds to gradient
Occurrence in wetlands dominated by tidal and
wind-generated water-level fluctuations

Kettles, potholes, vernal
pools, Carolina bays,
groundwater slope wetlands
Blanket bogs, tussock tundra

Riparian wetlands along
rivers, streams
Estuarine tidal wetlands,
lacustrine fringes subject to
winds, waves, and seiches

Ombrotrophic bogs, pocosins

Fens, groundwater slope
wetlands

Alluvial swamps, tidal
wetlands, montane
streamside wetlands

Usually depressional
wetlands, bogs (annual),
prairie potholes (multiyear)

Usually riverine wetlands

Usually fringe wetlands

Source: Brinson (1993)

of depressional wetlands to unidirectional flows found in riverine wetlands to
bidirectional flows of tidal and high-energy lacustrine wetland systems. Uni- and
bidirectional surface flows range widely in energy from hardly perceptible movement
to strong erosive currents. Combined with the geographic setting, hydrodynamics
can result in a range of different wetland types.

This classification system was designed to be independent of plant communities,
because it depends on the geomorphic and hydrologic properties of the wedands. In
practice, however, vegetation often provides important clues to the hydrogeomorphic
forces at work. Because most modern classification systems developed for inventory
purposes have some basis in hydrogeomorphology, their classes often give important
clues as to function.

Wetland Classifications 469

Rating Wetlands

Since the introduction of legislation and regulations aimed at wetland conser-
vation, especially in the United States, there has been considerable interest in
classifying wetlands for their value to society in order to simplify the issuance of
permits for wetlands activities; that is, high-value wetlands presumably would
receive more protection than low-value wetlands. Also, when wetland loss is
being mitigated, it is importantto compare the functions are being lost and those
that are gained with the mitigation strategy. Almost every state in the United
States has procedures for rating wetlands. Three examples are described here.

1. Washington

The state of Washington was one of the first places where such a wetland
“rating” was developed. Its system, with the state divided into two hydrogeo-
morphic areas, western and eastern, established a method for including the
state’s wetlands in one of four categories based on their sensitivity to dis-
turbance, their significance, their rarity, our ability to replace them, and the
functions they provide (Hruby, 2004).

Category I. Unique, rare, relatively irreplaceable, or high level of valuable
functions

Category II. Wetlands that are difficult though not impossible to replace
Category III. Moderate level of functions

Category IV. Lowest level of valuable functions and often heavily disturbed

The rating scores were based on water quality and hydrologic and
habitat functions (Hruby, 2004). Effective 2015, the state of Washington
revised its 2004 rating system by changing some of the rating questions.
See http://www.ecy.wa.gov/programs/sea/wetlands/ratingsystems/2014
updates.html

2. Ohio

Ohio developed a similar wetland assessment technique based on earlier ver-
sions of the Washington model. The Ohio Rapid Assessment Method (ORAM)
for wetlands attempts to determine the function of wetlands in three cate-
gories, put in reverse order to the Washington system (Mack, 2001):

Category 1. Wetlands that provide minimal functions or habitat
Category 2. Wetlands that provide moderate functions and are dominated by
native species

Category 3. Wetlands that provide superior functions, have high levels of
diversity, and are habitat for rare and endangered species

470 Chapter 13 Wetland Classification

A quantitative ranking system on a O-to-100 scale, based on field and
remote-sensed data, is at the heart of this system. Six metrics to determine
the wetland score include:

1. Wetland area

2. Upland buffers and surrounding land use

3. Hydrology

4. Habitat alteration and development

5. Special wetlands

6. Plant communities

Details are provided in Mack (2001), which is available at www.epa.state
.oh.us/dsw/401/ecology.aspx.

3. Florida

The state of Florida, which has more remaining wetlands than any other state
except Alaska and has one of the highest rates of human population increases
of any state, has major challenges of balancing human development, wetland
protection, and mitigation of wetland loss (see Chapter 18: “Wetland Creation
and Restoration" for discussion on mitigating wetland losses). In an attempt
to quantify and compare wetland losses with gains resulting from mitigation
of those losses, the state developed a state-wide wetland assessment
method that is called a Uniform Mitigation Assessment Method (UMAM).
It became effective in 2004. UNAM “is designed to assess any type of
impact and mitigation, including the preservation, enhancement, restoration,
and creation of wetlands, as well as the evaluation and use of mitigation
banks” (http://sfrc.ufl.edu/ecohydrology/UMAM_Training_Manual_ppt.pdf).
For the quantitative section, wetlands (to compare current conditions and
expected mitigation conditions, for example) are evaluated in the following
three categories on a scale of 0 to 10:

1. Location and landscape support — ^the ecological context in which the
wetland system functions;

2. Water environment — ^the hydrologic condition including degree to
which it is altered and impaired; and

3. Community structure (vegetation and/or benthic/sessile
communities) — vegetation measurements include plant cover,
species lists, invasive exotics, plant conditions, and topographic
features; benthic and sessile community structure is evaluated where
submerged benthic communities are present such as oyster reefs,
corals, and soft-bottom systems such as riverine systems.

Wetland Remote Sensing and Inventory 471

When mitigation of wetland loss is being proposed, the comparison of
the estimated indices from the above three factors are adjusted using a
preservation factor, a factor for time lag, and a factor that incorporates the
anticipated risk. Details of this wetland assessment method are at www.dep
■state.fl.us/water/wetlands/mitigation/umam. A useful training manual for
UMAM can be found at http://sfrc.ufl.edu/ecohydrology/UMAM_Training_
Manual_ppt.pdf.

Wetland Remote Sensing and Inventory

One of the major objectives of wetland classification is to be able to inventory the
location, extent, and type of wetlands in a region of concern. An inventory can be
made of a small watershed, a political unit such as a county or parish, an entire state
or province, or an entire nation. Whatever the size of the area to be surveyed, the
inventory must be based on some previously defined classification and should be con-
structed to meet the needs of specific users of information on wetlands. Generally,
inventories require not only information about the types and extent of wetlands but
also documentation of their geographic locations and boundaries. To accomplish this
goal, remote platforms — aircraft and/or satellites — produce imagery — ^photographs or
digital information that can be used to make images. This imagery must be interpreted
to identify the locations, boundaries, and types of wetlands on the image.

Remote-Sensing Platform

In the early days of wetland classification, wetlands were mapped from surveyors’
records and from boats. Later, interpretation of low- altitude aerial photographs com-
bined with verification in the field made the process both faster and more accurate.
High-altitude imagery from aircraft such as that from the U-2 were available in the
past, but satellite imagery is rapidly becoming the norm. The commonly used satellite
systems — Landsat Multispectral Scanner (MSS), Landsat Thematic Mapper (TM), and
Systeme Pour I’Observation de la Terre (SPOT) — now have been joined by a num-
ber of high-resolution environmental satellites (Table 13.5). These remote platforms
are an effective way of gathering data for large-scale wedand surveys. Satellites offer
repeated coverage that allow seasonal monitoring of wedands as well as providing data
on the surrounding landscape readily translatable to a geographic informadon system
(GIS) format (Ozesmi and Bauer, 2002). The choice of which platform to use depends
on the resoludon required, the area to be covered, and the cost of the data collection.
Low-aldtude aircraft surveys offer reladvely inexpensive and fairly effecdve ways to
survey small areas. High-aldtude aircraft offer much greater coverage in each image
(photograph) and may be less expensive per unit area than low-aldtude aircraft when
costs of photo interpretation are included. The limitations of satellite remote sens-
ing include the lack of ability to separate different wetland types or to even separate
wedands from upland forests or agricultural land.

472 Chapter 13 Wetland Classification

Table 13.5 Examples of spatial resolution, revisiting time, and aititude of recent satellites availabie for
wetiand inventories and studies

IKONOS

QuickBird

OrbView-3

WorldView-1

GeoEye-1

WorldView-2

Sponsor

Space Imaging

Digital Globe

Orbimage

Digital Globe

GeoEye

Digital Globe

Launch Date

Sept 1999

Oct 2001

June 2003

Sept 2007

Sept 2008

Oct 2009

Spatial resolution, m

1.0

0.61

1.0

0.5

0.41

0.5

(panchromatic)

Spatial resolution, m

4.0

2.44

4.0

n/a

1.65

(muitispectral)

Swath width, km

11.3

16.5

17.6

15.2

16.4

Revisit time, days

2.3-S.4

1-3.5

1.5-3

1. 7-3.8

2. 1-8.3

1.1-2. 7

Orbitai altitude, km

681

450

470

496

681

770

Source: Klimas, 2013

Orbiting satellites have been providing data for Earth resources classification since
the launching of the first of the Landsat satellites in 1972. Today, several highly effec-
tive satellites appropriate for wetland inventories and management orbit the Earth
(Klimas, 2013). One problem with early satellites was the poor resolution (Landsat
has a resolution of 30 m). Today’s satellites (Table 13.5) can resolve features ofEarth’s
surface to as little as 1 m, but high resolution is not an unmixed blessing. It requires the
ability to transmit and process enormous amounts of data, because the data generated
for a given surface area quadruples every time the resolution doubles.

Remote-Sensing Imagery

In addition to choosing the remote-sensing platform, the wedand scientist or man-
ager has the choice of several types of imagery from different types of sensors. Color
photography and color-infrared photography were popular for many years for wedand
inventories from aircraft (see Shuman and Ambrose, 2003), although black-and-white
photography has been used with some success. Color-infrared film and now digital
imagery provide good definidon of plant communides and is the film of choice. Satel-
lites, and some aircraft, gather digital data in one or more electromagnedc spectral
bands. For example, most of the satellites listed in Table 13.5 have panchromatic and
multispectral capability in several color bands, including infrared. The U.S. Nadonal
Wedands Inventory (see box) relies to some extent on imagery interpretation, but
computerized interpretation of satellite imagery is used more and more for wetland
mapping, especially on agricultural landscapes.

The U.S. National Wetlands Inventory

The U.S. National Wetlands Inventory (NWI) is a good example of a major
wetlands mapping project and illustrates some of the problems encoun-
tered in any mapping enterprise. The Cowardin et al. (1979) classification

Wetland Remote Sensing and Inventory 473

scheme has provided the basic mapping units for the NWI being carried
out by the U.S. Fish and Wildlife Service for several decades. The Service
announced as of May 2014 that all of the wetlands of the entire lower
48 states of the United States, Hawaii, and dependent territories as well
as 35 percent of Alaska are now digitally mapped (www.fws.gov/wetlands/
Documents/Completion-of-National-Wetlands-Database-News-Release.pdf),
and users can now access wetlands data and maps through an online wetland
mapping service.

For the NWI, aerial photography at scales ranging from 1:60,000 to
1:130,000 was the primary source of data, with color-infrared photography
providing the best delineation of wetlands (Wilen and Pywell, 1981; Tiner and
Wilen, 1983). In the 1970s, maps were created mostly from 1:80, 000-scale
black-and-white photography. It is now supplemented with Environmental Sys-
tems Research Institute satellite imagery, which provides resolution of Im
or better for the United States. Photointerpretation and field reconnaissance
are then used to define wetland boundaries according to the wetland

South Naples Bay,
Naples FL

Everglades Wetland
Research Park

PF04/3A « Palustrine
Forested Needle-Leaved
Evergreen/Broad-leaved
Evergreen Temporarily
Flooded

E2SS3/EM1P-Estuarine
Intertidal Shrub Scrub/
Emergent Persistent
Irregularly Flooded

User Remarks;

includes location of Everglades Wetland Resesrch Park

Figure 13.5 Sample of map of wetlands in Naples, Florida, created from the
U.S. National Wetlands Inventory at www.fws.gov/wetlands/index.html, showing
two examples of a classification notation, one for an estuarine intertidal shrub
scrub (= mangrove swamp) and the other for a palustrine forested needle-leaved
(= cypress swamp). Map also shows location of Everglades Wetland Research Park
(http://fgcu.edu/swamp) at the Naples Botanical Garden.

474 Chapter 13 Wetland Classification

classification system. The information is summarized on base maps using
an alphanumeric system based on the U.S. Fish and Wildlife Classification
system (Cowardin et al., 1979).

Nowadays wetland maps can be crafted at any scale by users at the
NWI Inventory site at www.fws.gov/wetlands/index.html using their “Wetland
Mapper” system (see Fig. 13.5). In Figure 13.5, two wetland indicators are
highlighted; one indicates an expansive mangrove swamp in an estuarine sys-
tem and the other shows a cypress swamp/hardwood swamp combination in
the palustrine system.

Recommended Readings

Brinson, M. M. 1993. A Hydrojjeomorphic Classification for Wetlands. Wetlands
Research Program Technical Report WRP-DE-4. Vicksburg, MS: U.S. Army
Corps of Engineers Waterways Experiment Station.

Cowardin, L. M., V. Carter, F. C. Golet, and E. T. LaRoe. 1979. Classification of
Wetlands and Deepwater Habitats of the United States. Washington, DC: U.S.
Fish and Wildlife Service, FWS/OBS-79/31.

Tiner, R. W. 1999. Wetland Indicators: A Guide to Wetland Identification, Delineation,
Classification, and Mapping. Boca Raton, EL: CRC Press.

U.S. National Wetland Inventory web page: www.fws.gov/wetlands/index.html.

References

Brinson, M. M. 1993. A Hydropfeomorphic Classification for Wetlands. Wedands
Research Program Technical Report WRP-DE-4, U.S. Army Corps of Engineers
Waterways Experiment Station, Vicksburg, MS.

Cowardin, L. M., V Carter, F. C. Golet, and E. T. LaRoe. 1979. Classification of
Wetlands and Deepwater Habitats of the United States. Washington, DG: U.S.
Fish and Wildlife Service, FWS/OBS-79/31.

Davis, G. A. 1907. Peat: Essays on its origin, uses, and distribution in Michigan. In
Report of the State Board Geoloptical Survey Michigan for 1906, pp. 95-395.
Hruby, T. 2004. Washington state wetland rating system for western Washington —
Revised. Publication # 04-06-025. Washington State Department of Ecology,
Olympia, 113 pp. + appen.

Klimas, V. 2013. Using remote sensing to select and monitor wetland restoration sites:
An overview. Journal of Coastal Research 29: 958-970.

Kulczynski, S. 1949. Peat bogs of Polesie. Acad. Pol. Sci. Mem., Set. B, No. 15.
356 pp.

References 475

Mack, J. J. 2001. Ohio Rapid Assessment Method for Wetlands, v. 5. User’s Manual
and Scoring Forms, Ohio EPA Technical Report WET/2001-1. Ohio Environ-
mental Protection Agency, Division of Surface Water/Wetland Ecology Unit,
Columbus, Ohio, 66 pp. + forms.

Martin, A. C., N. Hutchkiss, F. M. Uhler, and W. S. Bourn. 1953. Classification of
Wetlands of the United States. Special Science Report — Wildlife 20, U.S. Fish and
Wildlife Service, Washington, DC. 14 pp.

Moore, P. D., and D. J. Bellamy. 1974. Peatlands. Springer-Verlag, New York. 221 pp.

Odum, H. T., B. J. Copeland, and E. A. McMahan, eds. 1974. Coastal Ecolo£iical
Systems of the United States. Conservation Foundation, Washington, DC. 4 vols.

Ozesmi, S. L., and M. E. Bauer. 2002. Satellite remote sensing of wetiands. Wetlands
Ecology and Management 10: 381-402.

Potonie, R. 1908. Aufbau und Vegetation der Moore Norddeutschlands. En^lers
botanische jahrbiicher 90. Leipzig, Germany.

Shaw, S. P, and C. G. Fredine. 1956. Wetlands of the United States, their extent, and
their value for waterfowl and other wildlife. Gircular 39, U.S. Fish and Wildlife
Service, U.S. Department of Interior, Washington, DG. 67 pp.

Shuman, G. S., and R. F. Ambrose. 2003. A comparison of remote sensing and
ground-based methods for monitoring wetland restoration success. Restoration
Ecology 11: 325-333.

Tiner, R. W., and B. O. Wilen. 1983. U.S. Fish and Wildlife Service National Wetlands
inventory project. Unpublished report, U.S. Fish and WildUfe Service, Washing-
ton, DG. 19 pp.

Warner, B. G., and G. D. A. Rubec, eds. 1997. The Ganadian Wedand Glassification
System. National Wetlands Working Group, Wedands Research Gentre, Univer-
sity of Waterloo, Ontario.

Weber, G. A. 1907. Aufbau und Vegetadon der Moore Norddutschlands. Beibl. Bot.
Jahrbiichern. 90: 19-34.

Wilen, B. O., and H. R. Pywell. 1981. The National Wetlands Inventory. Paper pre-
sented at In-Place Resource Inventories: Principles and Practices — A Nadonal
Workshop, Orono, ME, August 9-14. 10 pp.

Chapter 14

Human Impacts and Management
of Wetlands

Wetland impacts have included both wetland alteration and wetland
destruction. In earlier times, wetland drainapfe was considered the only policy
for manapfin^ wetlands. The most common alterations of wetlands have been
draininpi, dred^in^, and filling of wetlands; modification of the hydrolopfic
regime; highway construction; mining and mineral extraction; and water
pollution. Peat resources, estimated to be 1.9 trillion tons in the world, are
harvested in many countries as a source of fuel and horticultural materials.
Wetlands can also be managed close to their natural state for certain objectives,
such as fish and wildlife enhancement, agricultural and aquaculture
production, water quality improvement, and flood control. Management of
wetlands for coastal protection has now taken on more significance with
potential sea-level increases.

The concept of wetland management has had different meanings at different times
to different disciplines and in different parts of the world. Until the middle of the
twentieth century, the term wetland management usually meant wedand drainage to
many policy makers, except for a few resource managers who maintained wedands
for hunting, fishing, and waterfowl/wildlife protecdon. Landowners were encouraged
through government programs to tile and drain wedands to make the land suitable for
agriculture and other uses. Dredging for navigation and filling for land development
destroyed countless coastal and inland wetlands.

Until the last quarter of the twentieth century, there was litde understanding
of and concern for the inherent values of wedands except by those who recognized
wedands as wildlife habitats, pardcularly for waterfowl. A whole science of “marsh
management” developed in the middle part of the twendeth century around the idea
of maintaining specific hydrologic condidons to opdmize fish or waterfowl popu-
ladons. Only since the mid-1970s have other values, such as flood control, coastal

477

478 Chapter 14 Human Impacts and Management of Wetlands

protection, and water quality enhancement, been recognized. It has taken disasters
such as the 1993 Upper Mississippi River Basin flooding, the 2004 Indian Ocean
tsunami, and the 2005 Hurricane Katrina disaster in New Orleans to cause societies
to focus on the potential lives that could be saved and property damage minimized if
wetland buffer systems were provided at our land-water margins.

Today, the management of wetlands usually means setting several objectives,
depending on the priorities of the wetland managers, current environmental regula-
tions, and wishes of a myriad of stakeholders who are usually involved. In some cases,
objectives such as preventing pollution from reaching wetlands and using wetlands
as sites of water quality improvement can be conflicting. Many floodplain wetlands
are now managed and zoned to minimize human encroachment and maximize
floodwater retention. Coastal wetlands are now included in coastal zone protection
programs for storm protection and as sanctuaries and subsidies for estuarine fauna. In
the meantime, wetlands continue to be altered or destroyed throughout the world by
drainage. Ailing, conversion to agriculture, water pollution, and mineral extraction.

We are thankful to have witnessed a slowing of the destruction rate of wetlands,
even since we wrote the first edition of this Wetland textbook (Mitsch and Gosselink,
1986), at least in the United States. We are not as certain that destruction of the
world’s wetlands is being slowed, but we are aware that there is a much greater inter-
national appreciation of wetlands than before. Vigilance is required, however, to make
sure that wetland values continue to be protected. Wetland conservation and even wet-
land restoration and creation (see Chapter 18: “Wetland Creation and Restoration,”
for details on this type of wetland management) have accelerated, particularly in the
developed world over the past 40 years. But there are few if any regulations or restric-
tions on wetland destruction or pollution in developing parts of the world. This may
be the next frontier of wetland protection.

Early History of Wetland Management

The early history of wetland management, a history that still influences many people
today, was driven by the misconception that wetlands were wastelands that should be
avoided or, if possible, drained and filled. Throughout the world, as long as there have
been humans, there has been hydrologic alteration of the landscape. As summarized
by Joe Larson and Jon Kusler (1979): “For most of recorded history, wetlands were
regarded as wastelands if not bogs of treachery, mires of despair, homes of pests, and
refuges for outlaw and rebel. A good wetland was a drained wetland free of this mixture
of dubious social factors.”

In the United States, this opinion of wetlands and shallow- water environments led
to the destruction of more than half of the total wetlands in the lower 48 states over
a 200-year period. In New Zealand, settlement by Europeans that began in earnest in
the mid-I800s contributed significantly to a 90 percent loss of wetlands in a relatively
short time. Preliminary estimates suggest that, over human history, about half of the
world’s wetlands have been lost (see Chapter 3: “Wetlands of the World”).

Early History of Wetland Management 479

Table 14.1 Human actions that cause direct wetland losses and degradation^

Freshwater

Lakes/Littoral

Swamp

Cause

Estuaries

Floodplains

Marshes

Zone

Peatlands

Forest

Agriculture, forestry, mosquito control

drainage

Stream channelization and dredging;

flood control

Filling — solid-waste disposal; roads;

development

Conversion to aquaculture/mariculture

Dikes, dams, seawall, levee

construction

Water pollution — urban and agricultural

Mining of wetlands of peat and other

materials

Groundwater withdrawal

‘'xx = common and important cause of wetland loss and degradation.
x = present but not a major cause of wetland loss and degradation.

Blank Indicates that effect is generally not present except in exceptional situations.
Source'. Dugan (1993)

With over 70 percent of the world’s population living on or near coastlines, coastal
wetlands have long been destroyed through a combination of excessive harvesting,
hydrologic modification and seawall construction, coastal development, pollution, and
other human activities. Likewise, inland wedands have been continually affected, par-
ticularly through hydrologic modification and agricultural and urban development.
Human activities, such as agriculture, forestry, stream channelization, aquaculture,
dam, dike, and seawall construction, mining, water pollution, and groundwater with-
drawal, all had impacts, some severe, on wetlands (Table 14.1). Wetlands are degraded
and destroyed indirectly as well through alternation of sediment patterns in rivers,
hydrologic alteration, highway construction, and land subsidence (Table 14.2). A
third possibility is the loss of wedands from natural causes (Table 14.3 ), although wet-
lands are normally resilient and can recover from natural events. For example, many
coastal wetlands that were devastated by the 2004 Indian Ocean tsunami or the 2005
hurricane that destroyed much of New Orleans have long since recovered.

The propensity in the East was not to drain valuable wedands endrely, as has been
done in the West, but to work within the aquadc landscape, albeit in a heavily managed
way. Dugan’s (1993) interesdng comparison between hydraulic civilizations (Euro-
pean in origin), which controlled water flow through the use of dikes, dams, pumps,
and drainage rile, and aquatic civilizations (Asian in origin), which better adapted
to their surroundings of water-abundant floodplains and deltas, is an interesdng way
to view humans’ use of wetlands. The former approach of controlling nature rather
than working it is becoming more dominant around the world today; that is why we
continue to find such high losses of wetlands worldwide.

480 Chapter 14 Human Impacts and Management of Wetlands

Table 14.2 Human activities that indirectly cause wetland losses and degradation^

Freshwater

Lakes/Littoral

Swamp

Cause

Estuaries

Floodplains

Marshes

Zone

Peatlands Forest

Sediment retention by dams and

other structures

Hydrologic alteration by roads,

canals, etc.

Land subsidence due to groundwater.

resource extraction, and river
alternations

®xx = common and important cause of wetland loss and degradation.
x= present but not a major cause of wetiand loss and degradation.

Blank indicates that effect is generally not present except in exceptional situations
Source: Dugan (1993)

Table 14.3 Natural events that cause wetland losses and degradation^

Cause

Estuaries

Floodplains

Freshwater

Marshes

Lakes/Littoral

Zone

Peatlands

Swamp

Forest

Subsidence

Sea-level rise

Drought

Hurricanes, tsunamis,

and other storms

Erosion

Biotic effects

®xx = common and important cause of wetland loss and degradation.
x= present but not a major cause of wetland loss and degradation.

Blank indicates that effect is generally not present except in exceptional situations
Source: Dugan (1993)

Wetland Drainage History in the United States

Had not politics intervened, George Washington may have succeeded in draining the
Great Dismal Swamp in Virginia in the mid-eighteenth century (see Chapter 3 ) instead
of leading a new nation. Draining swamps and other wetlands was an acceptable and
even desired practice from the time Europeans first settled in North America. In the
United States, public laws actually encouraged wetland drainage. Congress passed the
Swamp Land Act of 1849, which granted to Louisiana the control of all swamplands
and overflow lands in the state for the general purpose of controlling floods in the
Mississippi River basin. In the following year, the act was extended to the states of
Alabama, Arkansas, California, Llorida, Illinois, Indiana, Iowa, Michigan, Mississippi,
Missouri, Ohio, and Wisconsin. Minnesota and Oregon were added in 1860. The
act was designed to decrease federal involvement in flood control and drainage by

Wetland Drainage History in the United States 481

transferring federally owned wetlands to the states, leaving to them the initiative of
“reclaiming” wetlands through activities such as levee construction and drainage.

By 1954, an estimated 26 million ha of land had been ceded to those 15 states
tor reclamation. Ironically, although the federal government passed the Swamp Land
Act to get out of the flood control business, the states sold those lands to individuals
for pennies per acre, and the private owners subsequently successfully lobbied both
national and state governments to protect these lands from floods. Further, govern-
ments are now paying enormous sums to buy the same lands back for conservation
purposes. Although current government policies are generally in direct opposition to
the Swamp Land Act and it is now disregarded, the act cast the initial wedand policy
of the U.S. government in the direction of wedand elimination.

Other acdons led to the rapid decline of the nation’s wetlands. An estimated 23
million ha of wet farmland, including some wedands, were drained under the U.S.
Department of Agriculture’s Agricultural Conservation Program between 1940 and
1977. An esdmated 18.6 million ha of land, much of it wedands, was drained in seven
states in the upper Mississippi River basin alone. Some of the wedand drainage acdv-
ity was hastened by projects of groups such as the Depression-era Works Progress
Administradon, the Soil Conservadon Service, and other federal agencies. Coastal
marshes were eliminated or drained and ditched for intercoastal transportadon, res-
idendal developments, mosquito control, and even for salt marsh hay producdon.
Interior wetlands were converted primarily to provide land for urban development,
road construction, and agriculture.

Go South, Young Man?

Typical of the prevalent attitude toward wetlands in the mid-twentieth century
is the following quote by Norgress (1947) discussing the “value" of Louisiana
cypress swamps:

With 1,628,915 acres of cutover cypress swamp lands in Louisiana at the
present time, what use to make of these lands so that the Ideal cypress
areas will make a return on the Investment for the landowner Is a serious
problem of the future ....

The lumbermen are rapidly awakening to the fact that In cutting the
timber from their land they have taken the first step toward putting it in
position to perform its true function — agriculture ....

It requires only a visit Into this swamp territory to overcome such
prejudices that reclamation is impracticable. Millions of dollars are being
put Into good roads. Everywhere one sees dredge boats eating their way
through the soil, making channels for drainage.

After harvesting the cypress timber crop, the Louisiana lumbermen are
at last realizing that In reaping the crop sown by Nature ages ago, they have
left a heritage to posterity of an asset of permanent value and
service — land, the true basis for wealth.

482 Chapter 14 Human Impacts and Management of Wetlands

The day of the pioneer cypress lumberman is gone, but we need today
in Louisiana another type of pioneer — the pioneer who can help bring under
cultivation the enormous areas of cypress cutover lands suitable for
agriculture. It is important to Louisiana, to the South, and the Nation as a
whole, that this be done. Would that there were some latter-day Horace
Greeleys to cry, in clarion tones, to the young farmers of today, “Go South,
young man; go South!”

As an example of state action leading to wetland drainage, Illinois passed the Illi-
nois Drainage Levee Act and the Farm Drainage Act in 1879, which allowed counties
to organize into drainage districts to consolidate financial resources. This action accel-
erated draining to the point that 30 percent of Illinois and Indiana and 20 percent of
Iowa and Ohio are now under some form of drainage, and almost all of the original
wetlands in these states (80 to 90 percent) have been destroyed. Chapter 3 described
two very large wetlands in this region of the United States — the Great Kankakee Marsh
in Indiana and the Black Swamp in Ohio — that essentially no longer exist. Drainage
was absolute there.

Wetland Alteration

In a sense, wetland alteration or destruction is an extreme form of wetland manage-
ment. One model of wetland alteration (Fig. 14. 1) assumes that three main factors
influence wetland ecosystem health: water level, nutrient status, and natural distur-
bances. Through human activity, the modification of any one of these factors can lead
to wedand alteration, either directly or indirectly. For example, a wetland can be dis-
turbed through decreased water levels, as in draining and filling, or through increased
water levels, as in downstream drainage impediments. Nutrient status can be affected
through upstream flood control that decreases the frequency of nutrient inputs or
through increased nutrient loading from agricultural areas.

The most common alterations of wetlands have been (I) draining, dredging,
and filling of wetlands; (2) modification of the hydrologic regime; (3) highway con-
struction; (4) mining and mineral extraction; and (5) water pollution. These wetland
modifications are described in more detail next.

Wetland Conversion: Draining, Dredging, and Fiiling

The major cause of wetland loss around the world continues to be conversion to agri-
cultural use. Drainage for farms in the United States progressed at an average rate
of 490,000 ha/yr over much of the twentieth century (slope of line in Fig. 14. 2a).
Less drainage occurred during the Great Depression of the 1930s and World War II
years. This conversion was particularly significant in the vast midwestern United States

Wetland Alteration 483

Figure 14.1 Model of human-induced impacts on wetlands, including effects on water level,
nutrient status, and natural disturbance. By either increasing or decreasing any one of these
factors, wetlands can be altered. (After Keddy, 1983)

“breadbasket,” which has provided the bulk of the grain produced on the continent
(Fig. 14.2b). Some of the world’s richest farming is in the former wetlands of Ohio,
Indiana, Illinois, Iowa, and southern Minnesota. When drained and cultivated, the
fertile soils of the prairie pothole marshes and east Texas playas also produce excel-
lent crops. With ditching and modern farm equipment, it has been possible to farm
former wetlands routinely (Fig. 14.3). The modern farm equipment of today and
mass-produced reels of plastic drainage pipe also make it possible to drain much more
area per day than was ever possible with earlier equipment and the use of clay tiles.

Some of the most rapid wetland losses have occurred in the bottomland hardwood
forests of the lower Mississippi River alluvial floodplain (Fig. 14.4). As populations
increased along the river, the floodplain was channeled and leveed so that it could
be drained and inhabited. Since colonial times, the floodplain has provided excellent
cropland, especially for cotton and sugarcane. Cultivation, however, was restricted to
the relatively high elevation of the natural river levees, which flooded regularly after
spring rains and upstream snowmelts but drained rapidly enough to enable farmers to
plant their crops. Because spring floods naturally fertilized river levees, they required
no additional fertilizers to grow productive crops. One of the results of drainage and
flood protection is the additional cost of fertilization. The lower parts of the flood-
plain, which are too wet to cultivate, were left as forests but harvested for timber.
As pressure for additional cropland increased, these agriculturally marginal forests
were clear-cut at an unprecedented rate. This was feasible, in part, because of the

484 Chapter 14 Human Impacts and Management of Wetlands

Figure 14.2 Artificially drained land in the United States: (a) Tt'end from 1900 to 1980; (b)
extent and location of drainage through mid-1980s. Each dot represents 8,000 ha (20,000
acres), and the total area drained is 43 million ha. ((a) After Gosselink and Maltby, 1990; (b)
after Dahl, 1990)

development of soybean varieties that mature rapidly enough to be planted in June
or even early July, after severe flooding has passed. Often the land thus reclaimed
was subsequently incorporated behind flood control levees, where it was kept dry by
pumps. Clear-cutting of bottomland forests is still proceeding. Most of the available

Wetland Alteration

485

Figure 14.3 Modern drainage machinery such as that illustrated in these photos is abie to
drain dozens of hectares per day: (a) Detail of the drainage machinery; (b) results of about
1 minute of drainage, showing new ditch and plastic pipe installed. (Photos by W. J. Mitsch)

wetland has been converted in Arkansas and Tennessee; Mississippi and Louisiana are
experiencing large losses.

Along the nation’s coasts, especially the East and West Coasts, the major cause of
wetland loss is draining and filling for urban and industrial development or wetland loss
due to subsidence. Compared to land converted to agricultural use, the area involved is
rather small. Nevertheless, in some coastal states, notably California, almost all coastal
wetlands have been lost. The rate of coastal wetland loss from 1954 to 1974 was
closely tied to population density. This finding underscores two facts: (1) Two-thirds
of the world’s population lives along coasts; and (2) population density puts great
pressure on coastal wetlands as sites for expansion. The most rapid development of
coastal wetlands occurred after World War 11. In particular, several large airports were
built in coastal marshes. Since the passage of federal legislation controlling wetland
development, the rate of conversion has slowed.

486 Chapter 14 Human Impacts and Management of Wetlands

(ymmaixy,

|FOF?EST0(^O'mER ORIGINAL TYP«5i CURRENT (19db
CROPLAND. PASTURE. OR URBW.

LOWER MISSISSIPPI RIVER BASIN

BEFORE 1600

Figure 14.4 Historical and current distribution of bottomland wetland ferests in the Missis-
sippi River floodplain. (After The Nature Conservancy, 1992)

Hydrologic Modifications

Ditching, draining, and levee building are hydrologic modifications of wetlands specif-
ically designed to dry them out. Other hydrologic modifications destroy or change
the character of thousands of hectares of wetlands annually. Usually these hydrologic
changes were made for some purpose that had nothing to do with wedands; wetland
destruction is an inadvertent result. Canals, ditches, and levees are created for three
primary purposes:

1 . Flood control. Most of the canals and levees associated with wedands are for
flood control. The canals have been designed to carry fioodwaters off the
adjacent uplands as rapidly as possible. Normal drainage through wedands is
slow surface sheet flow; straight, deep canals are more efficient. Ditching
marshes and swamps to drain them for mosquito control or biomass
harvesdng is a special case designed to lower water levels in the wedands.
Along most of the nation’s major rivers are systems of levees constructed to
prevent overbank flooding of the adjacent floodplain. The U.S. Army Corps
of Engineers built most of these levees after Congress passed flood control

Wetland Alteration 487

legislation following the disastrous floods of the 1920s and 1930s. (For a
fascinating account of the great flood of 1927, the disruption it caused, and
the social and political reverberations that led to flood control legislation, see
Barry, 1997.) These levees, by separating the river from its floodplain,
isolated wetlands so that they could be drained expeditiously. For example,
along the lower Mississippi River, the construction of levees created a
demand from farmers for additional floodplain drainage. The sequence of
response and demand was so predictable that farmers bought and cleared
floodplain forests in anticipation of the next round of flood control projects.

2. Navigation and transportation. Navigation canals tend to be larger than
drainage canals. They traverse wetlands primarily to provide water
transportation access to ports and to improve transport among ports. For
example, the Intracoastal Waterway was dredged through hundreds of miles of
wetlands in the northern Gulf Coast. In addition, when highways were built
across wetlands, fill material for the roadbed was often obtained by dredging
soil from along the right-of-way, thus forming a canal parallel to the highway.

3. Industrial activity. Many canals are dredged to obtain access to sites within a
wetland to sink an oil well, build a surface mine, or other kinds of
development. Usually pipelines that traverse wetlands are laid in canals that
are not backfilled.

The result of all of these activities can be a wetland crisscrossed with canals, such
as in the immense coastal wetlands of the northern Gulf Coast. These canals modify
wetlands in many ecological ways by changing normal hydrologic patterns. Straight,
deep canals in shallow bays, lakes, and marshes capture flow, depriving the natural
channels of water. Canals are hydrologically efficient, allowing the more rapid runoff
of freshwater than the normal shallow, sinuous channels do. As a result, water lev-
els fluctuate more rapidly than they do in unmodified marshes, and minimum levels
are lowered, drying the marshes. In addition, when deep, straight channels connect
low-salinity areas to high-salinity zones, as with many large navigation channels, tidal
water, with its salt, intrudes farther upstream, changing freshwater wedands to brack-
ish. In extreme cases, salt-intolerant vegetation is killed and is not replaced before the
marsh erodes into a shallow lake. On the Louisiana coast, the natural subsidence rate is
high; wedands go through a natural cycle of growth followed by decay to open bodies
of water. There, canals accelerate the subsidence rate by depriving wetlands of natural
sediment and nutrient subsidies.

Highway Construction

Highway construcdon can have a major effect on the hydrologic condidons of wet-
lands. Although few definidve studies have been able to document the extent of
wedand damage caused by highways, the major effects of highways are alteradon of the
hydrologic regime, sediment loading, and direct wedand removal. In general, wedands
are more sensidve to highway construcdon than uplands are, pardcularly through the

488 Chapter 14 Human Impacts and Management of Wetlands

disruption of hydrologic conditions. Many early studies (Clewell et al., 1976; Evink,
1980; and Adamus, 1983) found that highway construction led to negative effects on
wetlands through hydrologic isolation. Other than solar energy, the most important
driving forces for wetlands are hydrologic, including tides, gradient currents (e.g.,
streamflow), runoff, and groundwater flow. The importance of protecting the hydro-
logic regime during highway construction is based on the contention presented in
Chapter 4, “Wetland Hydrology,” that the hydrology of wetlands is the most impor-
tant determinant of a wetland’s structure and function.

Peat Mining

World resources of peat, principally in peatlands in the Northern Hemisphere, are
estimated to be 1.9 X lO'^^ t (trillion metric tons), of which countries that comprise
the former Soviet Union have about 770 X 10^ t (billion tons) and Canada about
510 X 10^ t. In the United States, deposits of peat occur in most states, with esti-
mated resources of about 310 X 10^ t, or about 16 percent of the world total. Surface
peat mining has been a common activity in several European countries, particularly
Ireland and countries in eastern and northeastern Europe, since the eighteenth cen-
tury (Fig. 14.5). These countries account for almost 75 percent of peat mining in the
world; some of this peat is still used as a fuel for electric power production. For cen-
turies but no longer, turf (dried-out peat) was used for home hearing in Ireland (see
Chapter 1: “Wetlands: Human Use and Science”). Peat projection in the world was
estimated to be 25.5 X 10^ t/yr in 2012, down from previous years and about the
same amount of peat production as 14 years prior (Table 14.4).

Figure 14.5 Peat mining near Tartu, Estonia. Peat is burned in power piant shown with
smokestack in background. (From J. S. Aber; printed with permission)

Wetland Alteration 489

Table 14.4 World peat production by country for 1998 and 2012 in metric tons (t) per year

1998 Peat Production (x 10^ t/yr)

2012 Peat Production (x 10^ t/yr)

Country'^

Fuel

Horticulture

Total

Fuel

Horticulture

Total

Finland

7,000

400

7,400

4,000

760

4,760

Ireland

4,500

300

4,800

1,452

500

1,952

Russia

3,000

1,300

Germany

180

2,800

2,980

3,048

Canada

1,127

973

Sweden

800

250

1,050

1,880

1,420

3,300

Ukraine

1,000

735

Estonia

1,000

360

567

927

United States

676

488

United Kingdom

500

Latvia

450

13,800

Belarus

300

250

3,000

3,250

Netherlands

300

Moldova

475

Denmark

205

130

France

200

Poland

200

736

Lithuania

195

371

386

Spain

Flungary

Norway

440

Australia

n/a

Argentina

Burundi

Turkey

150

Rwanda

Grand total

16,800

6,900

25,500

11,200

9,200

25,500

®ln addition to the countries listed, Austria, Chile, Iceland, Italy, and Romania produced negligible amounts of
peat.

Source: Jasinski (1999) and United States Geological Survey (2013)

Peat produced in the United States was about 448,000 1 (metric tons) per year in
2012, ranking the country twelfth in total production in the world. Since its inception,
peat mining in North America has been primarily for horticultural and agricultural
applications. The fibrous structure and porosity of peat promote a combination of
water retention and drainage, which makes it useful for applications such as potting
soils, lawn and garden soil amendments, and turf maintenance on golf courses. Peat is
also used as a filtering medium to remove toxic materials and pathogens from waste-
water, sewage effluent, and stormwater. It is generally classified as reed-sedge peat,
whereas the imports from Canada typically are a weakly decomposed Sphagnum peat,
which has a higher market value per ton. Approximately 95 percent of domestic peat is
sold for horticulture/agriculture usage, including, in order of importance, general soil
improvement, potting soils, earthworm culture, the nursery business, and golf course
maintenance and construction.

490 Chapter 14 Human Impacts and Management of Wetlands

Figure 14.6 The impact ef ceai surface mining on wetlands and the possible use of wetlands
in reciamation of ceai surface mines for wildlife enhancement and control of mine drainage.

Mineral and Water Extraction

Surface mining activity for materials other than peat often affects major wetlands
regions. Phosphate mining in central Florida is carried out over 120,000 ha and
has had a significant impact on wetlands in the region (Brown, 2005). Thousands
of hectares of wetlands may have been lost in central Florida because of this activity
alone, although the reclamation of phosphate-mined sites for wetlands is now a com-
mon practice. H. T. Odum et al. (1981) argued that “managed ecological succession”
on mined sites could be an economical alternative to current expensive reclamation
techniques involving massive earth moving and reclamation planting.

Surface mining of coal has also affected wetlands in some parts of the United States
(Brooks et ah, 1985). Forty-six thousand hectares of wetlands in western Kentucky
in the early 1980s, mostly bottomland hardwood forests, were or could have been
affected by surface coal mining. The recognition of the potential benefits of including
wetlands as part of the reclamation of coal mines has not been as widespread as one
would have expected (Fig. 14.6), because of strict interpretation of measures regulat-
ing the return of the land to its original contours and because of liability questions.
This is in contrast to the widespread acceptance of the reclamation of wetlands on
phosphorus mine sites in Florida.

In some parts of the country, the withdrawal of water from aquifers or minerals
from deep mines has resulted in accelerated subsidence rates that are lowering the ele-
vations of marshes and built-up areas alike, sometimes dramatically. Land subsidence,
which can also result in the creation of lakes and wetlands, is a geologically common
phenomenon in Florida. Often, when excessive amounts of water are removed from
karst deposits, underground cave-ins occur, causing surface slumpage. Some believe
that the cypress domes in north-central Florida are an indirect result of a similar natural

Wetland Management by Objective 491

process, whereby fissure and dissolutions of underground limestone cause slight sur-
face slumpage and subsequent wetland development.

Water Pollution

Wedands are altered by pollutants Ifom upstream or local runoff and, in turn, change
the quality of the water flowing out of them. The ability of wetlands to cleanse water
has received much attention in research and development and is discussed elsewhere
in this book. The effects of polluted water on wetlands have received less attention,
although water quality standards for wetlands have now been established in several
regions of United States.

Species composition may also change with eutrophication of wetlands. For
example, increased agricultural runoff, laden with phosphorus, is believed to have
caused a spread of Typha domin^ensis in conservation areas that are part of the
original Everglades in Florida (Fig. 14.7). This, in turn, has increased fears that the
phosphorus will eventually lead to invasion of Typha in the Everglades National Park,
replacing the natural sawgrass [Cladium jamaicense) (see Case Study 1, Chapter 18:
“Wetland Creation and Restoration,” and Case Study 3, Chapter 19: “Wetlands and
Water Quality”).

When metals, oils or other toxic organic compounds are the pollutants, effects on
the wetland can be dramatic such as in the 2010 Deepwater Horizon Gulf of Mex-
ico oil spill, where about 1800 km of coastal wetlands were affected (NRC, 2013).
Another case of pollution in wetlands occurred two decades before when sulfates
were discharged into a forested wetland in Florida (J. Richardson et al., 1983). Acid
drainage from active and abandoned coal mines has been shown to affect wetlands
seriously. In a study of wetlands adjacent to coal surface mining in western Kentucky,
Mitsch et al. (1983a, 1983b, 1983c) described the extensive ecological damage that
could occur where waters with low pH and high iron and sulfur were discharged from
the mines into or through wetlands.

In one of the most publicized and dramatic cases of water pollution of a wet-
land, selenium from farm runoff contaminated marshes in Kesterson National Wildlife
Refuge in California’s San Joaquin Valley (Ohlendorf et al., 1986, 1990; Presser and
Ohlendorf, 1987; T. Harris, 1991). The selenium contamination led to excessive death
and deformities of wildlife and to eventual “closing” of the contaminated marsh in the
mid-1980s, amid much controversy.

Wetland Management by Objective

Wetlands are managed for environmental protection, for recreation and aesthetics, and
for the production of renewable resources. Twelve specific goals of wetland manage-
ment are applicable today:

1 . Maintain water quality.

2. Reduce erosion.

492 Chapter 14 Human Impacts and Management of Wetlands

Everglades
Agricultural Area

H Everglades agricultural area
l^fia-dominated marsh
Mixed Typha-Oadium marsh
I I Oadium marsh and aquatic sloughs

Figure 14.7 Water Conservation Area 2A (44,700 ha) in the south Florida Everglades, show-
ing the area that has received high-nutrient surface overfiow from agricuitural land drainage
since the 1960s. Excess nutrients frem the Everglades Agricultural Area te the northwest
have caused the spread of Typha domingensis and the loss of Cladium jamaicense over the
8,000-ha area shaded. (After Koch and Reddy, 1992)

3. Protect from floods and storm damage.

4. Provide a natural system to process airborne pollutants.

5. Provide a buffer between urban residential and industrial segments to
ameliorate climate and physical impact such as noise.

6. Maintain a gene pool of marsh plants and provide examples of complete
natural communities.

7. Provide aesthetic and psychological support for human beings.

Wetland Management by Objective 493

8. Produce wildlife.

9. Control insect populations.

10. Provide habitats for fish spawning and other food organisms.

11. Produce food, fiber, and fodder (e.g., timber, cranberries, cattails for fiber).

12. Expedite scientific inquiry.

One management approach is to fence in a wetland to preserve it. Although sim-
ple, this is an act of conservation of a valuable natural ecosystem involving no substan-
tive changes in management practices. Often, however, management has one or more
specific objectives that require positive manipulation of the environment. Efforts to
maximize one objective may be incompatible with the attainment of others, although,
in recent years, most management objectives have been broadly stated to enhance mul-
tiple objectives. Multipurpose management generally focuses on system-level support
rather than individual species. This has often been achieved indirecdy through plant
species manipulation, because plants provide food and cover for the animals. In the
management of many small wetland areas in proximity, the use of different practices or
staggered management cycles, so that the different areas are not all treated the same
way at the same time, not only increases the diversity of the larger landscape but also
attracts wildlife.

Waterfowl and Wildlife Management

The best wetland management practices are those that enhance the natural processes
of the wedand ecosystem involved. One way to accomplish this is to maintain con-
didons as close as possible to the natural hydrology of the wetland, including hydro-
logic connections with adjacent rivers, lakes, and estuaries. Unfortunately, this cannot
easily be accomplished in wedands managed for wildlife; the vagaries of nature, espe-
cially in hydrologic conditions, make planning difficult. Elence, marsh management
for wildlife, particularly waterfowl, has often meant water-level manipuladon. Dikes
(impoundments), weirs (solid structures in marsh outflows that maintain a minimum
water level), control gates, and pumps control water level. In general, the results of
the management acdvity depend on how well the water-level control is maintained,
and control depends on the local rainfall and on the sophisdcation of the control
structures. For example, weirs provide the poorest control; all they do is maintain a
minimum water level. Pumps provide posidve control of drainage or flooding depth
at the desired dme; and the management objectives usually can be met, although the
cost is much higher than fixed weirs.

Baldassarre and Bolen (2006) summarize several of the wetland management
techniques used for waterfowl and water birds. They conclude that the practices fall
into two general categories: (1) natural management thsx takes advantage of natu-
ral attributes of wetlands, such as seed banks, plant succession, water-level fluctua-
tions, and herbivory; and (2) artificial management thzt includes practices such as
planting, ditching, and island building. One of the most frequently used manage-
ment techniques — ^perhaps a combination of the two above categories — is a water-level

494 Chapter 14 Human Impacts and Management of Wetlands

moist soil perennial

annuals emergents

floating submerged
aquatics aquatics

open

water

vegetation characteristics

annuals

annuals and
perennials

perennials

summer water level

mudflat by
mid-June

15 cm

>30 cm

seed production (annuals)

maximum

fair

low

plant species diversity

fair

maximum

fair

plant density

maximum

good

low

resident wildlife use

fair

maximum

good

aquatic invertebrate
and fish abundance

fair

maximum

good

migratory wildlife use

maximum

good

fair

invasion potential of
undesirable plants

high

low

muskrat production

low

good

from R.W. Kroll

Figure 14.8 Generalizations of water-level management for vegetation, wildlife use, and
other characteristics as practiced on impounded marshes near Lake Erie in northern Ohio.
(Ftom Roy Kroll, unpublished illustratien, previded with permissien)

drawdown. Drawdowns are carried out to recycle nutrients from otherwise undecom-
posed organic matter, to allow for “moist-soil management” to enhance vegetation
regeneration from the wetland seed bank, and sometimes to manage for a diversity of
macroinvertebrate (important source of protein for ducks) communities. Quite often,
trade-offs occur during water-level manipulations.

To illustrate the trade-offs in wetland management for wildlife enhancement,
some generalizations about water-level manipulation of Lake Erie (Ohio) coastal
marshes are shown in Figure 14.8. Maximum migratory wildlife use of the marshes
occurs in moist-soil conditions, but these conditions are also the best for the invasion
of potentially undesirable plants and are generally least favorable for the overall
abundance and diversity of resident plant and animal populations. Shallow- water
(called hemi conditions by marsh managers; around 15 cm depth in summer) usually
result in the highest plant species diversity and greatest fish and resident wildlife use

Wetland Management by Objective 495

but less migratory wildlife. Deepwater conditions (>30cm) offer the least potential
for both annual emergent plants and invading, undesirable plants, and desirable
migratory waterfowl use is only fair in deep water. Kroll et al. (1997) and Gottgens
et al. (1998) point out that because the landward advance of marshes during high
lake water times is restricted by human development, and because carp (Cyprinus
carpio) are present in the lake, long-term above-average water levels probably mean
that removal of dikes along the Great Lakes would lead to an irreversible loss of
wetland vegetation fringing the lakes. Mitsch et al. (2001 ) found that only 25 percent
of the existing marshes encompassed by dikes would have had the right conditions to
be emergent marshes more than 50 percent of the time during the twentieth century.

The set of management recommendations by Weller (1978) for prairie pothole
marshes in the north-central United States and south-central Ganada is another
example of multipurpose wildlife enhancement. Those recommendations mimic the
natural cycle of marshes in the middle of North America. Although they may seem
drastic, they are entirely natural in their consequences. In sequence, the six practices
are:

1 . When a pothole is in the open stage and there is litde emergent vegetation,
the cycle should be initiated by a spring drawdown. This stimulates the
germination of seedlings on the exposed mud surfaces.

2. A slow increase in water level after the drawdown maintains the growth of
flood -tolerant seedlings without shading them out in turbid water. Shallowly
flooded areas attract dabbling ducks during the winter.

3. The drawdown cycle should be repeated for a second year to establish a good
stand of emergent plants.

4. Low water levels should be maintained for several more seasons to encourage
the growth of perennial emergent plants such as Typha.

5. Maintaining stable, moderate water depths for several years promotes the
growth of rooted submerged perennial aquatic plants and associated benthic
fauna that make excellent food for waterfowl. During that period, the
emergent vegetation will gradually die out and will be replaced by shallow
ponds. When that occurs, the cycle can be initiated again, as described in
step 1.

6. Different wetland areas maintained in staggered cycles provide all stages of
the marsh cycle at once, maximizing habitat diversity.

Weller (1994) made the distinction between a complete drawdown of water levels,
as described before — a management option when vegetation is completely lost because
of high water levels, herbivory, winter kill, or plant disease — and a partial drawdown,
which can be implemented when vegetation is reduced but not eliminated or when
wildlife use has declined but not disappeared.

Wildlife management in coastal salt marshes such as those found in Louisiana uses
a similar strategy, although the short-term cycle is not as pronounced there. Draw-
downs to encourage the growth of seedlings and perennials preferred by ducks are

496 Chapter 14 Human Impacts and Management of Wetlands

common practices, as is fall and winter flooding to attract dabbling ducks. As it hap-
pens, there is general agreement that stabilizing water levels is not good management,
even though our society seems to feel intuitively that stability is a good thing. Wetlands
thrive on cycles, especially flooding cycles, and practices that dampen these cycles also
reduce wildlife productivity. Although the management practices described previously
enhance waterfowl production, they are generally deleterious for wetland-dependent
fisheries in coastal wetlands because free access between the wetlands and the adja-
cent estuary is restricted; the wetlands’ role in regulating water quality is also often
underutilized.

There is a tendency to want to control all external variables when we manage
wetlands by objectives. This management tendency, although understandable, is par-
ticularly strong when herbivores, such as geese, nutria, beavers, or muskrats, “invade”
a managed wetland. These animals can be discouraged and/or trapped to keep their
influence on vegetation at a minimum, but one has to remember that these animals
are not invaders at all but are simply coming to a habitat that is generally well suited
to their needs. In the ecosystem context, these animals are often nature’s “ecosystem
engineers” and provide many functions that, in the long term, may enhance marshes.
Beavers cause water-level manipulations just as humans do. Muskrats and geese remove
large areas of vegetation but open up the system to allow for other vegetation to come
into the wetland.

Whether management of ecosystem managers is a wise strategy is a complex issue.
In coastal Louisiana, for example, muskrats and especially nutria (a South American
immigrant) can “eat-out” extensive marsh areas, which do not recover because of
rising sea level and high marsh subsidence rates induced, in part, by human activities.
Trapping used to keep the rodent populations in check, but the worldwide slump
in fur sales no longer makes trapping profitable, and rodent populations are rapidly
escalating.

Baldassarre and Bolen (2006) presented seven general principles that provide a
useful set of rules for wetland managers. Many of the principles on wetland restoration
described in Chapter 18 mirror these wetland management principles:

1 . Protect wetland complexes that include a wide variety of wetland
hydroperiods and wetland sizes.

2. Protect small wetlands, as these wetlands are most vulnerable to being lost
along with their unique biota.

3. Consider all wetland-dependent wildlife when managing wetlands, not just
one or two species.

4. Protect large wetlands too for species that require large areas.

5. Recognize the importance of wetland complexes for species with complex
life-history requirements.

6. Recognize that protected sites often require direct management mtervention
to protect wildlife values.

7. Protect and restore upland habitats that are contiguous with wetlands.

Wetland Management by Objective 497

Agriculture and Aquaculture

When wetlands are drained for agricultural use, they no longer function as wetlands.
They are, as local farmers say, “fast lands” removed from the effects of periodic flood-
ing, and they grow terrestrial, flood -intolerant crops. Some use is made of more or
less undisturbed wetlands for agriculture, but it is minor. In New England, high salt
marshes were harvested for salt marsh hay {Spartina patens)^ which was considered
an excellent bedding and fodder for cattie. In fact, the proximity of fresh and salt
hay marshes was a major factor in selecting the sites for many towns in New England
before 1650. Subsequently, marshes were ditched to allow the intrusion of tides to
promote the growth of salt marsh hay, but the extent of this practice has not been
well documented. On parts of the coast of the Gulf of Mexico where marshes are firm
underfoot, they are still used extensively for catde grazing. To improve access, small
embankments or raised earthen paths are constructed in these marshes.

The ancient Mexican practice of marcmo is unique. In the freshwater wedands
of the northern coast of Mexico, small areas were cleared and planted in corn during
the dry season. These native varieties were tolerant enough to withstand considerable
flooding. After harvest (or apparently sometimes before harvest), the marshes were
naturally reflooded, and native grasses were reestablished until the next dry season.
This practice is no longer followed, but there has been some interest in reviving it.

On a global scale, the production of rice in managed wedands contributes a major
proporrion of the world’s food supply. There are approximately 1.3 million km^ of rice
paddies in the world (Chapter 3), of which almost 90 percent are in Asia. In North
America, especially in Minnesota, there are several commercial wild rice {Zizania
aquatica) operarions in wedands and several other locadons where Native American
tribes have harvested wild rice in natural marshes for centuries.

Aquaculture, the farming of fish and shellfish, which produced less than I million
tons per year in the early 1950s, now produces 70 million metric tons, or almost half
of the annual worldwide total fish and shellfish harvest of 160 X 10*’ t (million metric
tons)/yr (Fig. 14.9). The Food and Agriculture Organizadon of the United Nadons
(http://www.fao.org/fishery/topic/I3540/en) esdmates that “to maintain the cur-
rent level of per capita consumpdon, global aquaculture producdon will need to reach
80 million ton by 2050.” Most of this aquaculture producdon occurs in Asia, with
China by far the largest producer. The United States is the major consumer of aqua-
culture products but accounts for only a small percentage of worldwide producdon,
mostly salmon and crayfish.

Fish farming practices vary. The most environmentally benign approach, similar
to the Mexican marceno described previously, intercrops shellfish with a grain crop,
usually rice. Typical is crayfish farming in the United States and Indian shrimp aqua-
culture in rotadon with rice. The pracdce is described for crayfish in the southern
United States. Crayfish are an edible delicacy in the southern United States and in
many foreign countries. They live in burrows in shallow flooded areas, such as swamp
forests and rice fields, emerging with their young early in the year to forage for food.
The young grow to edible size within a few weeks and are harvested in the spring.
When floodwaters retreat, the crayfish construct burrows, where they remain undl the

498 Chapter 14 Human Impacts and Management of Wetlands

1950 1960 1970 1980 1990 2000 2010

Figure 14.9 World’s fisheries harvest, 1950 to 2010, showing increased contribution of
aquaculture to overall production. (From http://en.wikipedia.org/wiki/Seafood)

next winter flood. In crayfish farms, this natural cycle is enhanced by controlling water
levels. An area of swamp forest is impounded; it is flooded deep during the winter and
spring and drained during the summer. This cycle is ideal for crayfish, which thrive.
Fish predators are controlled within the impoundments to improve the harvest. The
hydrologic cycle is also favorable for forest trees. It simulates the hydrologic cycle of
a bottomland hardwood forest; forest tree productivity is high, and seedling recruit-
ment is good because of the summer drawdown. Species composition tends toward
species typical of bottomland hardwoods.

Some rice farmers have also found that they can take advantage of the annual
flooding cycle typically used to grow rice to combine rice and crayfish production.
Rice fields are drained during the summer and fall, when the rice crop matures and
is harvested. Then the fields are reflooded, allowing crayfish to emerge from their
burrows in the rice field embankments and forage on the vegetation remaining after
the rice harvest. The crayfish harvest ends when the fields are replanted with rice. When
this rotation is practiced, extreme care has to be exercised in the use of pesticides.

The most intensive aquaculture techniques control all aspects of production. Wet-
lands, salt flats, mangrove forests, and even high-quality farmland are dredged to form
ponds in which water levels are controlled by pumps. “Seed” organisms, the young
postlarvae, are raised in separate hatcheries. The young organisms are fed in the ponds
on synthetic diets, often composed of fish from commercial fisheries. Water quality is
monitored, the ponds are aerated, and, in the most sophisticated operations, wastes
are treated. Yields from this kind of operation can be several metric tons per hectare
per crop, and in tropical areas two crops per year are expected.

Recommended Readings

Connolly, K. D., S. M. Johnson, and D. R. Williams. 2005. Wetlands Law and Policy.
Chicago: American Bar Association.

National Research Council. 1995. Wetlands: Characteristics and Boundaries. Wash-
ington, DC: National Academy Press.

National Research Council. 2001. Compensating for Wetland Losses under the Clean
Water Act . Washington, DC: National Academy Press.

Ramsar Convention on Wetlands Web site: www.ramsar.org

References

Downing, D. M., C. Winer, and L. D. Wood, 2003. Navigating through Clean Water
Act Jurisdiction: A Legal Review. Wetlands 23: 475M:93.

Environmental Defense Fund and World Wildlife Fund. 1992. How Wet Is a Wetlandl
Lhe Impact of the Proposed Revisions to the Federal Wetlands Delineation Manual.
Environmental Defense Fund and World Wildlife Fund, Washington, DC. 175 pp.

Gardner, R. C. 2014. Contemplating Koontz: Ten takeaways. National Wetlands
Newsletter 36(2): 9-11.

Goldman-Carter, J. 2014. U.S. Supreme Court to water resource managers: Be careful
what you ask for. National Wetlands Newsletter 36(2): 12-14.

Murphy, J. E. 2006. Rapanos v. United States: Wading through murky waters.
National Wetlands Newsletter 28(5): 1.

National Research Council. 1995. Wetlands: Characteristics and Boundaries. National
Academy Press, Washington, DC. 306 pp.

National Research Council. 2001. Compensating for Wetland Losses under the Clean
Water Act . National Academy Press, Washington, DC. 158 pp.

National Wetlands Policy Forum. 1988. Protecting America’s Wetlands: An Action
Apienda. Conservation Foundation, Washington, DC. 69 pp.

Ramsar Convention. 2011. Ramsar’s Liquid Assets, 40 years of the Convention on
Wetlands. Ramsar Wetland Convention, Gland, Switzerland, 36 pp.

Runyon, L. C. 1993. The Lucas Court Case and Land-Use Planning. National Confer-
ence of State Legislators, Denver, CO, Supplement to State Legislatures, vol. 1,
no. 10 (March).

U.S. Army Corps of Engineers. 1987. Corps of Engineers Wedands Delineation man-
ual. Technical Report Y-87-1. U.S. Army Corps of Engineers Waterways Experi-
ment Station, Vicksburg, MS. 100 pp. and appendices.

P a r t V

Ecosystem Services

Chapter 16

Wetland Ecosystem Services

Wetlands provide many services and commodities to humanity. Using the
Millennium Ecosystem Assessment terminology, provisioning services include
harvesting of wetland-dependent fish, shellfish, fur-bearing animals, waterfowl,
timber, and peat. Regulating ecosystem services from wetlands include
moderating the effects of floods, improve water quality, protect coastlines from
storms, hurricanes, and tsunamis, climate regulation, and aquifer recharge.
Cultural ecosystem services have aesthetic and heritage values and subsistence of
ancient and sustainable cultures. Valuation techniques include nonmonetary
scaling and weighting approaches for comparing different wetlands or
different management options for the same wetland and common-denominator
approaches that reduce the various values to some common term, such as dollars,
embodied energy, or emergy. These common-denominator methodologies can
include willingness to pay, replacement value, energy analysis, and emergy
analysis. None of these approaches is without problems, and no universal
agreement about their use has been reached. But when compared to other
ecosystems or uses of the landscape, sustainable ecosystem services of wetlands are
often among the highest of any ecosystems.

The terms value and services impose an anthropocentric orientation on a discussion
of wetlands. In ordinary parlance, the words connote something worthy, desirable,
or useful to humans. The reasons that wetlands are often legally protected have to
do with their value to society, not with the abstruse ecological processes that occur in
wetlands; this is the sense in which the words value and services 's.cc used in this chapter.
Perceived values arise from the functional ecological processes described in previous
chapters but are determined also by human perceptions, the location of a particular
wetland, the human population pressures on it, and the extent of the resource.

527

528 Chapter 16 Wetland Ecosystem Services

Regional wetlands are integral parts of larger landscapes — drainage basins and/or
estuaries. Their functions and their values to people in these landscapes depend on
both their extent and their location. Thus, the value of a forested wetland varies. If
it lies along a river, it probably plays a greater functional role in stream water quality
and downstream flooding than if it was isolated from the stream. If situated at the
headwaters of a stream, a wetland functions differendy from a wedand located near
the stream’s mouth. The fauna it supports depend on the size of the wedand reladve
to the home range of the animal. Thus, to some extent, each wedand is ecologically
unique. This complicates the measurement of its value.

Wetland Ecosystem Services

A series of publicadons in the mid-2000s referred to as the Millennium Ecosystem
Assessment (2005) brought focus to the importance of natural ecosystems by poindng
out four main findings with regard to humans and the planet’s ecosystems:

1. Humans have changed the planets ecosystems over the last 50 years of the
twentieth century more than any comparable period in human history.

2. The changes that have been made to ecosystems have contributed to
substandal gains in well being and economic development, but at the cost of
losing many ecosystem services.

3. The degradadon is expected to grow significandy worse in the first half of the
twenty-first century.

4. Reversing this degradation of ecosystems will involve significant changes in
policies, pracdces, and insdtudons that are not yet in place.

The term ecosystem services replaced ecosystem values, a concept used in earlier edidons
of this textbook since its first edidon in 1986. In those previous edidons, we divided
the values of wedands into three hierarchical levels familiar to ecologists: population,
ecosystem, and global. Population values include those related to providing habitats
for animals harvested for pelts, waterfowl and other hunted and watched birds, fish
and shellfish producdon, timber and peat harvesdng, and support of endangered and
threatened species. Ecosystem values included water quality improvement, storm and
flood midgation, aquifer recharge, and even sustenance of human cultures. Global
values included maintaining water and air quality influences on a much broader scale
than that of the ecosystem level, especially in regional and global cycles of nitrogen,
sulfur, and carbon.

The current paradigm of ecosystem services as organized by the Millennium
Ecosystem Assessment (2005) also divided ecosystem services into three categories
but with a division related to human well-being (Fig. 16. 1).

I . Provisioning ecosystem services include products obtained from ecosystems,
such as food, water, rimber, fiber, or generic resources.

Wetland Ecosystem Services 529

ECOSYSTEM SERVICES

Provisioning

FOOD

PHESHWATfc..

■•/OOO AND FI8EH

Supporting

NUTRIENT CYCUNG
.-04L K)HMATION
PHlUAHYPnOOUCIlON

Regulating

'.UP4ATE HEGULATIOrj
ROOD REGULATION
DISEASE REfX"J^1fON
//A1ER PUfllF>CAl<3N

Cultural

At>THETl'.

^i'lh^lRJAL

EfXKiAItONA.

RECREATKMAi

UFE ON EARTH - BIODIVERSITY

CONSTITUENTS OF WELL-BEING

Security
i^HSONAL SAFETY
SECURE RESOURCE ACCESS
SECURITf FROM DISASTERS

Freedom
of choice
and action

OPPORTUNITY TO BE
ABLE TO ACHIEVE
WHATANINDMOUAL
VALUES DOING
ANDBGNG

Basic material
for good life

ADEQUATE irVEUHOOOS
SUFFICIENT Nurnmous food
SHELTER

ACCESSTOGOO06

Health

STRENGTH
FEELING WEU
ACCESS TO CLEAN AIR
ANDWATBI

Good social relations

SOCIAL COHESION
MUTUAL RESPECT
ABILITY TO Hap OTHERS

Source' Mlemurn Eooeystem Assesament

COLOR

Potential for mediation by
socioeconomic factors

WIDTH

Intensity of linkagea between ecoayetem
servicea and human welt-being

Low

Weak

Medium

■ ■ Medium

High

I I Strong

Figure 16.1 Illustration of provisioning, regulating, and cultural ecosystem services and
their connection to human weii-being. (From Miiiennium Ecosystem Assessment, 2005, copy-
right World Resources Institute, reprinted with permission)

2. Rejjulatinjf ecosystem services include air quality regulation, climate regulation,
water purification, disease regulation, pest regulation, pollination, and natural
hazard regulation.

3 . Cultural ecosystem services include benefits that people obtain from
ecosystems related to spiritual enrichment, recreation, ecotourism, aesthetics,
formal and informal education, inspiration, and cultural heritage.

This is the system we use in this current edition of Wetlands to describe wetland
ecosystem services.

Provisioning Ecosystem Services

Animals Harvested for Pelts

Fur-bearing mammals, and even alligators and crocodiles, are harvested for their
pelts throughout the world. In contrast to most other commercially important
wetland species, these animals typically have a limited range and spend their fives
within a short distance of their birthplaces. The most abundant fur-bearer historically

530 Chapter 16 Wetland Ecosystem Services

(b) (c)

Figure 16.2 Three fui^bearing animals feund in wetlands that have been historically har-
vested for their pelts: (a) muskrat {Ondatra zibethicus), (b) nutria {Myocastor coypus),
(c) beaver {Castor canadensis).

harvested in wetlands in the United States is the muskrat (Ondatra zibethicus).
Muskrats (Fig. 16.2a) are found in wetlands throughout the United States except,
strangely, the south Atlantic Coast. They prefer fresh inland marshes but along the
northern Gulf Coast are more abundant in brackish marshes. About 50 percent of
the nation’s harvest is from the Midwest and 25 percent from along the northern
Gulf of Mexico, mostly Louisiana. The nutria (Myocastor coypus)., an ecological analog
of the muskrat, is the next most abundant species. It is very much like a muskrat
but is larger and more vigorous (Fig. 16.2b). This species was imported from South
America to Louisiana and escaped from captivity in 1938, spreading rapidly through
the state’s coastal marshes. In the 1940s, the animal was promoted by state agents
for controlling aquatic weeds, particularly water hyacinth (Eichhornia crassipes). It is
now abundant in freshwater swamps and in coastal freshwater marshes, from which
it may have displaced muskrats to more brackish locations, and is spreading up the
coastal Atlantic states well beyond Louisiana. In order of decreasing abundance in
the United States, other harvested fur animals are beaver, mink, and otter. Beavers
( Castor canadensk. Fig. 16.2c) once had a population of 60 million in North America,
and were associated with a major colonial beaver-trapping industry, especially for

Wetland Ecosystem Services 531

European women’s fur and men’s and women’s hats and maintained largely by the
French in Canada and what is now the midwestern United States. Nowadays beavers
are associated with forested wetlands, especially in the Midwest. Minnesota harvests
a high percentage of the nation’s beaver catch in the United States. Beavers are
also now harvested for castoreum — a fungus exudate from the castor sacs of mature
beavers that is used for perfume, medicine, and a food additive.

Waterfowl and Other Birds

Birds, as our only remaining evolutionary link to the dinosaurs that once roamed
Earth, may have survived the dinosaur die-off precisely because of wetlands (Gib-
bons, 1997; Weller, 1999). Although not all current bird species require wedands as
their primary habitat, a great many do, and several are synonymous with wetlands
around the world (Figs. 16.3 and 16.4). Eighty percent of America’s breeding bird
population and more than 50 percent of the 800 species of protected migratory birds
rely on wetlands. Wetlands, which are probably known best for their waterfowl abun-
dance, also support a large and valuable recreational hunting industry. We use the term
industry because hunters spend large sums of money in the local economy for guns,
ammunition, hunting clothes, travel to hunting spots, food, and lodging.

Most of the birds hunted are hatched in marshes in the far North, sometimes
above the Arctic Circle, but are shot during their winter migrations to the south-
ern United States and Central America. There are exceptions — the wood duck {Aix
sponsa) breeds locally throughout the continent — but the generalization holds for
most species. Different groups of geese and ducks have different habitat preferences,
and these preferences change with the maturity of the duck and the season.

A broad diversity of wetland habitat types is important for waterfowl success. The
freshwater prairie potholes of North America are the primary breeding place for water-
fowl in North America. There, an estimated 50 to 80 percent of the continent’s main
game species are produced. Wood ducks prefer forested wetlands. During the winter,
diving ducks {Aythya spp. and Oxyum spp.) are found in brackish marshes, preferably
adjacent to fairly deep ponds and lakes. Dabbling ducks {Anas spy).) prefer freshwa-
ter marshes and often graze heavily in adjacent rice fields and in very shallow marsh
ponds. Gadwalls {Anas strepera) like shallow ponds with submerged vegetation.

The waterfowl value of wetlands such as the prairie pothole region of North Amer-
ica (see Chapter 3: “Wetlands of the World”) is unmistakable. When waterfowl census
data for the prairie pothole region over the 30-year period were compared to the
number of potholes flooded in May of each year, there was a clear positive corre-
lation, indicating the importance of wetland hydrology in the breeding success of
waterfowl. On average, there are almost 22 million waterfowl (dabbling and diving
ducks) in the region, dominated by the mallard. Generally, the duck population of
North America has shown a 10- to 20-year cycle of increase and decline, with low
points in the early 1960s and 1990s and highs in the mid-1950s, mid-1970s, and
late 1990s (Table 16.1). Populations of 9 of the 10 duck species listed in Table 16.1
were lower than historical averages after the dry years 1987 to 1991, while popula-
tions of 7 of the same 10 duck species were higher than historical averages after the

532 Chapter 16 Wetland Ecosystem Services

(b)

Figure 16.3 Two wetland waterfowl known around the world: (a) Mallard (Anas platyrhyn-
chos) and (b) Canada goose (Branta canadensis). (Photos courtesy of Alan and Elaine
Wilson)

wet years 1995 to 1998. Over that period, from dry period to wet period, the total
number of ducks increased by 60 percent. The trends of below-average populations
during dry periods and above-average populations during wet periods are particularly
apparent for mallards, green-winged and blue-winged teals, northern shovelers, and
canvasback. Climatic changes that influence the number of ponds from year to year in
the breeding grounds appear to be the major cause of year-to-year fluctuations.

Figure 16.4 Herons are consummate symbols of wetlands throughout the world. Different
species that dominate this wading niche in parts of the world include: (a) great blue heron
{Ardea herodias) from North America; (b) white-necked heron {Ardea cocol) from South Amer-
ica; (c) black-headed heron {Ardea melanocephala) from eastern Africa; (d) white-faced heron
(Ardea novaehollandiae) from Australia/New Zealand; (e) gray heron (Ardea cinerea) from
Europe and Africa. (Photograph (a) by T. Daniel, Ohio Department of Natural Resources; (b),
(c) by W. J. Mitsch; (d) by B. Harcourt, courtesy of New Zealand Department of Conservation;
(e) by P. Marion; reprinted by permission)

533

534 Chapter 16 Wetland Ecosystem Services

Table 16.1 Population estimates of the 10 most common species of breeding ducks and 4
species of goose in North America for a dry year (1991) and a wet year (1998) in the prairie
pothole region, with percentage change in 1991 and 1998 compared to 1995-1990 and
1955-1997 averages, respectively

Population (xl,000)

Percentage Change

Species

1991 (Dry Year)

1998 (Wet Year)

1991®

1998'’

All species

24,200

39,100

Mallard (Anas platyrhynchos)

5,353 ± 188

9,640 ± 302

-27

-f32

Gadwall (Anas strepera)

1,573 ± 94

3,742 ± 206

-f22

-f149

American wigeon (Anas americana)

2,328 ± 135

2,858 ± 145

-14

Green-winged teal (Anas crecca)

1,601 ± 88

2,087 ± 139

-4

-f16

Blue-winged teal (Anas discors)

3,779 ± 245

6,399 ± 332

-10

-f36

Northern shoveler (Anas clypeata)

1,663 ± 84

4,120 ± 194

-8

-f106

Northern pintail (Anas acuta)

1,794 ± 199

3,558 ± 194

-62

-36

Redhead (Aythya americana)

437 ± 37

918 ± 77

-26

-f48

Canvasback (Aythya valisneha)

463 ± 57

689 ± 57

-16

-f28

Scaup (Aythya spp.)

5,247 ± 333

4,122 ± 234

-7

-35

Average of 10 duck species
Canada goose (Branta canadensis)

3,750

4,683

-15

-f34

Snow goose (Chen caeruisecens)

2,440

3,776

White-fronted goose (Anser aibifrons)

492

941

Brant (Branta bernicia)

275

276

^Compared to average for 1955 to 1990.

‘’Compared to average for 1955 to 1997.

Source: U.S. Fish and Wildlife Service. Duck surveys on summer breeding grounds; goose surveys during
summer, fall, and winter.

Fish and Shellfish

A direct relationship between shrimp and fish harvests and wetland area has been
illustrated for many fisheries around the world, including marine, freshwater, and pond
raised (Fig. 16.5). Over 95 percent of the fish and shellfish species that are harvested
commercially in the United States are wetland dependent (Feierabend and Zelazny,
1987). The degree of dependence on wetlands varies widely with species and with the
type of wetland. Some important species are permanent residents; others are merely
transients that feed in wetlands when the opportunity arises. Some shallow wedands,
which may exhibit several other wetland values, may be virtually devoid of fish, whereas
other types of deepwater and coastal wedands may serve as important nursery and
feeding areas.

Virtually all of the freshwater species are dependent, to some degree, on wedands,
often spawning in marshes bordering lakes or in riparian forests during spring flood-
ing. These species are primarily recreadonal, although some small local commercial
fisheries exploit them. The saltwater species tend to spawn offshore, move into the
coastal marsh “nursery” during their juvenile stages, and then emigrate offshore as
they mature. They are often important for both commercial and recreadonal fisheries.
The menhaden is caught only commercially, but compeddon between commercial and

Wetland Ecosystem Services 535

Figure 16.5 Relationship between wetland area and fish harvests. The linear siope
describes the line of about 60 kg/ha yieid. (After Tkirner, 1982)

sport fishermen for shrimp, blue crab, oyster, catfish, sea trout, and striped bass can
be intensive and acrimonious. Anadromous fish probably use wetlands less than the
other two groups. However, young anadromous fish fry sometimes linger in estuaries
and adjacent marshes on their migrations to the ocean from the freshwater streams in
which they were spawned.

Analyses of fishery harvests from wetlands show the importance of recreational
fishing. Although the commercial harvest is usually much better documented, several
studies have shown that the recreational catch far outweighs the commercial catch
for certain species. Furthermore, the value to the economy of recreational fishing
is usually far greater than the value of the commercial catch, because sports fisher-
men spend more money per fish caught (they are less efficient) than their commercial
counterparts.

Timber and Other Vegetation Harvest

Wetlands often provide an abundance of building materials and foodstuffs for local
economies. Timber from forested wetlands was one of the staples of the economy of
southeastern United States. The antebellum homes of the South were often supported
by giant trusses of cypress trees harvested from nearby swamps. The Mississippi River
alluvial floodplain and the floodplains of rivers entering the South Atlantic are mostly
deciduous wetlands, whereas the forested wetlands along the northern tier of states
are primarily evergreen. The former are more extensive and potentially more valuable
commercially because of the much faster growth rates in the South.

536 Chapter 16 Wetland Ecosystem Services

In addition to the timber harvest, the production of herbaceous vegetation in
marshes is a potential source of energy, fiber, and other commodities. These prospects
have not been explored widely in North America but are viable options elsewhere.
For example, many commercial products are harvested from restored and natural salt
marshes and freshwater marshes in China. The productivity of many wetland species
(e.g., Spartina alterniflora (salt marsh cordgrass), Phra^mites australis (reed grass),
Typha spp. (cattail), Eichhornia crassipes (water hyacinth), Cyperus papyrus (papyrus)
is as great as our most vigorous agricultural crops.

Peat Harvesting

In addition to the annual production of living vegetation in wetlands, great reser-
voirs of buried peat exist around the world. Peat harvesting was described in detail
in Chapter 14: “Human Impacts and Management of Wetlands.” This buried peat
is a nonrenewable energy source that destroys the wetland habitat when it is mined.
In the United States and Canada, peat is mined primarily for horticultural peat pro-
duction, but in other parts of the world — for example, several republics of the former
Soviet Union and in Finland — ^it has been used as a fuel source for hundreds of years.
It is used to generate electricity, formed into briquettes for home use, and gasified or
liquefied to produce methanol and industrial fuels.

Endangered and Threatened Species

Wetland habitats are necessary for the survival of a disproportionately high percentage
of endangered and threatened species. Table 16.2 summarizes the statistics but imparts
no information about the particular species involved, their location, wetland habitat
requirements, degree ofwetland dependence, and factors contributing to their demise.
Although wetlands occupy only about 3 . 5 percent of the land area of the United States,
of the more than 200 animal species listed as endangered, about 50 percent depend
on wetlands for survival and viability. Almost one-third of native North American

Table 16.2 Threatened and endangered species associated with wetlands

Taxon

Number of
Species
Endangered

Number of
Species
Threatened

Percentage of
U.S. Total Threatened
or Endangered

Plants

Mammals

—

Birds

Reptiles

Amphibians

Mussels

—

Fish

Insects

Total

Source: Niering (1988)

Wetland Ecosystem Services 537

freshwater fish species are endangered, threatened, or of special concern. Almost all of
these were adversely affected by habitat loss. Sixty-three species of plants and 34 species
of animals that are considered endangered, threatened, or candidates for listing occupy
southern U.S. forested wetlands. Of these, amphibians and many reptiles are especially
linked to wetlands. In Florida, where the number of amphibian and reptile species is
about equal to the number of mammal and breeding bird species, 18 percent of all
amphibians and 35 percent of all reptiles are considered threatened or endangered or
their status is unknown (Harris and Gosselink, 1990).

The fate of one wetland-dependent endangered species is discussed here to illus-
trate the ecological complexity of species endangerment and also hope of recovery
of endangered species. Whooping cranes {Grus americcma) nest in wedands in the
Northwest Territories of Canada, in water 0.3 to 0.6 m deep, during the spring and
summer. In the fall, they migrate to the Aransas National Wildlife Refuge, Texas, stop-
ping off in riverine marshes along the migration route. In Texas, they winter in tidal
marshes. All three types of wedands are important for their survival. The decline in the
once -abundant species has been attributed both to hundng and to habitat loss. The
last whooping crane nest in the United States was seen in 1889. In 1941, the flock
consisted of 13 adults and 2 young. Since then, the flock has been gradually built up
to about 600 birds wild and capdve.

The American Alligator: From Endangered to Plentiful

The American alligator (Alligator mississippiensis; Fig. 16.6) represents a dra-
matic success story of the return from the edge of extinction to a healthy U.S.
population. Alligators are abundant in fresh and slightly brackish lakes and
streams and build nests in adjacent marshes and swamps in the southwestern
United States, especially in Florida and Louisiana. Alligators have an interest-
ing role in wetlands — they depend on them, and, in return, the character of
the wetland is shaped by the alligator, at least in the south Florida Everglades.
They are another example of an ecosystem engineer (See Chapter 7 : “Wetland
Vegetation and Succession.” As the annual dry season approaches, alligators
dig “gator holes." The material thrown out around the holes forms a berm high
enough to support trees and shrubs in an otherwise treeless prairie. The trees
provide cover and breeding grounds for insects, birds, turtles, and snakes. The
hole is a place where the alligator can wait out the dry period until the winter
rains. It also provides a refuge for dense populations offish and shellfish (up
to l,600/m2). These organisms, in turn, attract top carnivores, and so the
gator holes are sites of concentrated biological activity that may be important
for the survival of many species.

American alligator populations were reduced by hunters and poachers to
such low levels that the species was declared endangered in the 1970s.

538 Chapter 16 Wetland Ecosystem Services

Figure 16.6 The American aiiigator (Alligator mississippiensis) in Corkscrew Sanctu-
ary, Napies, Fiorida. (Photograph by W. J. Mitsch)

The species was threatened by severe hunting pressure, not by habitat loss.
When that pressure was removed, its numbers increased rapidly. The ani-
mal is now harvested under close regulation and grown commercially in both
Louisiana and Florida. About 250,000 alligators are harvested in the wild
and in farms annually in Louisiana, yet the population remains constant or
is slightly increasing. Alligator hunting and farming in Louisiana has increased
dramatically; it was worth $16 million in 1992 and $26 million in 2004 for
both wild and farm-raised animals.

In Florida, where limited hunting is permitted, the harvest in the wild and
on farms is considerably less than that in Louisiana, but the compatibility
of alligators and a rapidly increasing human population is constantly being
challenged. It is probably extraordinary that there have been fewer than 20
confirmed fatal alligator attacks on humans per recent 50 years in Florida,
given the high number of both alligators and people in the state.

In addition to the harvest of alligators for their meat, alligator skins from
both Florida and Louisiana are sold worldwide, particularly for high-end luxury
handbags, wallets, belts, and boots. Apparently with the increased interest in
wetlands and wildlife, the fashion world has gone reptile chic.

Wetland Ecosystem Services 539

Precipitation,
storms I

t '

Figure 16.7 The general effect ef wetlands on streamflow and stormwater runoff.

Regulating Ecosystem Services

Flood Mitigation

Chapter 4 dealt with the importance of hydrology in determining the character of
wetlands. In addition, wetlands influence regional hydrology. One way they do this is
to intercept storm runoff and to store storm waters, thereby changing sharp runoff
peaks to slower discharges over longer periods of time (Fig. 16.7). Because it is usually
the peak flows that produce flood damage, the effect of the wetland area is to reduce
the danger of flooding. Riverine wedands are especially valuable in this regard. In
a classic study on the Charles River in Massachusetts, the floodplain wetlands were
deemed so effective for flood control by the U.S. Army Corps of Engineers that it
purchased them rather than build expensive flood control structures to protect Boston
(U.S. Army Corps of Engineers, 1972). The study on which the Corps’ decision was
based demonstrated that if the 3,400 ha of wetlands in the Charles River basin were
drained and leveed off from the river, flood damages would increase by $17 million
per year.

Bottomland hardwood forests along the Mississippi River before European setde-
ment stored floodwater equivalent to about 60 days of river discharge. Storage capacity
has been reduced to only about 12 days as a result of leveeing the river and draining
the floodplain. The consequences — the confinement of the river to a narrow channel
and the loss of storage capacity — are major reasons that flooding is increasing along
the lower Mississippi River.

540 Chapter 16 Wetland Ecosystem Services

Novitzki ( 1985) analyzed the relationship between flood peaks and the percentage
of basin area in lakes and wetlands. In the Chesapeake Bay drainage basin, where the
wetland area was 4 percent, flood flow was only about 50 percent of that in basins
containing no wetland storage. However, in Wisconsin river basins that contained 40
percent lakes and wetlands, spring streamflow was as much as 140 percent of that in
basins that do not contain storage. This apparent anomaly is probably related to a
reduction in the proportion of precipitation that can inflltrate the soil and to a lack
of additional storage capacity in lakes and wetlands that are already at full capacity
during spring floods. Thus, the location of wetlands in the river basin can complicate
the response downstream. For example, detained water in a downstream wetland of
one tributary can combine with flows from another tributary to increase the flood
peak rather than to desynchronize flows.

Ogawa and Male (1983, 1986) used a hydrologic simulation model to investigate
the relationship between upstream wetland removal and downstream flooding. Their
study found that for rare floods — that is, those predicted to occur only once in 100
or more years — the increase in peak stream flow was signiflcant for all sizes of streams
when wetlands were removed. The authors concluded that the usefulness of wedands
in reducing downstream flooding increases with ( 1 ) an increase in wetland area, (2) the
distance that the wedand is downstream, ( 3 ) the size of the flood, (4) the closeness to an
upstream wedand, and (5) the lack of other upstream storage areas such as reservoirs.

Storm Abatement and Coastal Protection

Coastal wetlands absorb the first fury of ocean storms as they come ashore (Fig. 16.8).
Salt marshes and mangrove wetlands act as giant storm buffers (Barbier et ah, 2013;
Das and Crepin, 2013; Marois and Mitsch, 2015). This value can be seen in the
context of wetland conservadon versus development. Natural marshes and mangrove

vegetation protects sediments and
shallow depth from erosive forces
and reduces resuspension

Figure 16.8 The general protection that coastal wetlands provide to buffer coastlines from
tidal surges caused by hurricanes, typhoons, and tsunamis.

Wetland Ecosystem Services 541

forests, which sustain little permanent damage from these storms, can shelter inland
developed areas. Buildings and other structures on the coast are vulnerable to
storms, and hurricane and typhoon damage in the world is increasing almost every
year. Inevitably, the public pays much of the cost of this damage through taxes tor
public assistance, rebuilding public services such as roads and utilities, and federally
guaranteed insurance. Two coastal disasters in the first decade of the twenty-first
century poignandy illustrate in hindsight the value of coastal wetlands for coastal
protection (see boxes). In both of these cases, as the memory of the disaster fades,
there will be the tendency to go back to the ways things were done in the past.

Mangrove Swamps and the Indian Ocean Tsunami of December 2004

On December 26, 2004, an earthquake-caused tsunami produced unprece-
dented damage and loss of life (estimated that 230,000 were killed or miss-
ing) around the entire Indian Ocean. The earthquake activity center was off the
west coast of Sumatra, Indonesia, and so the greatest devastation occurred
in that region. This Boxing Day Tsunami has been called “one of the deadli-
est natural disasters in recorded history” (http://en.wikipedia.org/wiki/2004
Indian Ocean earthquake).

While no coastal defense system is capable of buffering areas that were
hit with a 10-m-high wall of water, it is clear that the destruction of mangrove
wetlands for shrimp farms and tourist meccas and the habitation of these
areas by humans are at least partially responsible for the carnage. The man-
grove swamps suffered significant temporary destruction as well, but they have
evolved to survive a violent seascape and certainly restored themselves. The
same cannot be said for human settlements that were built in areas of former
mangrove swamps.

One year prior to the Indian Ocean tsunami event, simulation models had
illustrated that a wide (100 m) belt of dense mangrove trees (referred to as a
“greenbelt”) could reduce a tsunami pressure flow by more than 90 percent
(Hiraishi and Harada, 2003). That information was not made public quickly,
and the Indian Ocean tsunami happened with little to no warning. In the five
countries hit hardest by the tsunami, at least 1.5 million ha of mangrove wet-
lands, or 26 percent of the mangrove cover, were destroyed between 1980
and 2000 (FAO, 2003; Check, 2005).

The protective role that mangrove wetlands provided during the Indian
Ocean tsunami was illustrated in hindsight for a region along the southeast
coastline in Tamil Nadu, India (Danielsen et al., 2005). In an area without
mangroves and coastal Casuarina plantations, a sand spit was totally removed
and parts of the local village was destroyed; there were “significantly less
damaged” areas where mangroves and plantations were present. Danielsen
et al. (2005) concluded that “conserving or replanting coastal mangroves and

542 Chapter 16 Wetland Ecosystem Services

greenbelts should buffer communities from future tsunami events.” There is
hope that such a tsunami disaster will never occur again, but conserving and
restoring mangrove swamps for coastal protection now has the attention of
all tropical and subtropical countries that face open oceans (See also Case
Study 6 in Chapter 18: “Wetland Creation and Restoration.”)

Hurricane Katrina of 2005 and New Orieans, Wetiand Wet Suit

Hurricane Katrina struck the Louisiana coastland and the city of New Orleans,
Louisiana, in late August 2005 with devastating results to lives and prop-
erty (Fig. 16.9). One of the reasons for the extensive destruction is the fact

INDEX

1 : AUGUST 23, 2005
2: AUGUST 26, 2005
3: AUGUST 28, 2005 7:15
4: AUGUST 29, 2005 14:45

lATEGORY 5
I^RRICANE

CATEGORY 2

iTroptca.' Depression
iTroptC^ S4orm
IC-ri(^>ory 1
ICale^ry 2 Humcane
[CalAgofy 3 Humcarte
Calegory 4 Humcaoe
i>l«9ory S Ht#ncane

Figure 16.9 Path of Hurricane Katrina across Florida, the Gulf of Mexico, Louisiana,
and Mississippi in August 2005. The hurricane crossed to the east of New Orleans on
August 29, 2005, and a tidal wave caused by the hurricane caused extensive damage
and loss of life in New Orleans and surrounding parishes. A more robust system of
coastal wetlands and barrier beaches, many of which have been lost in the last cen-
tury, would have provided more protection for the city. (From National Oceanic and
Atmospheric Administration)

Wetland Ecosystem Services 543

that New Orleans and Louisiana are losing their deltaic wetlands due to land
subsidence caused by natural and human effects. Studies over 50 years in
Louisiana led to the conclusion that “New Orleans was becoming a more vul-
nerable city with each passing year” (Costanza et al., 2006). The formerly
extensive salt marshes and other wetlands that used to surround New Orleans
could have provided some coastal protection from the 6-m storm surge that
overwhelmed the city’s levee system during Hurricane Katrina. But the wet-
lands have been lost at a rate of 65 km^ per year since the beginning of the
20th century, after 6,000 years of gradual land building. Almost 4,800 km^ of
coastal wetlands have been lost since the 1930s alone (Day et al., 2005).

Since marsh plants hold and accrete sediments (Cahoon et al., 1995),
often reduce sediment resuspension (Harter and Mitsch, 2003), and conse-
quently maintain shallow water depths, the presence of vegetation contributes
in two ways: (1) by actually decreasing surges and waves, and (2) by maintain-
ing the shallow depths that also accomplish the same. Because wetlands
indicate shallow water, the presence of wetland vegetation is also an “indi-
cator” of the degree to which New Orleans and other human settlements are
protected. While few experimental studies or modeling efforts have specifically
addressed the effect of coastal marshes on storm surges, anecdotal data
accumulated after Hurricane Andrew in 1992 in Louisiana suggested that the
storm surge from that hurricane was reduced about 4.7 cm per km of marsh
that it traveled over (Louisiana Coastal Wetlands Conservation Task Force and
Wetlands Conservation and Restoration Authority, 1998). Extrapolating from
this number, a storm tracking from the south of New Orleans through existing
coastal marshes could have its surge reduced by 3.7 m if it crossed 80 km
of marsh before reaching the city. Barbier et al. (2013) found that a 1 per-
cent increase in the wetland/open water ratio in the Louisiana delta could
decrease a storm surge by 8 to 11 percent, equivalent to a 1 m reduction in
storm surge per 9 to 13 km of additional wetlands. They translated this storm
surge reduction in to substantial reduction in property damage in southeastern
Louisiana. It is not inappropriate to refer to the disappearing marshes around
New Orleans as that city’s wet suit.

Climate Regulation

Wetlands may be significant factors in the global cycles of nitrogen, sulfur, and
carbon. The natural supply of ecologically useful nitrogen comes from the fixation
of atmospheric nitrogen gas (N2) by a small group of plants and microorganisms
that can convert it into organic form. Currendy, ammonia is manufactured from N2
for fertilizers, at more than double the rate of all natural fixation. Wedands may be
important in returning a part of this “excess” nitrogen to the atmosphere through
denitrihcadon. Denitrihcadon requires the proximity of an aerobic and a reducing

544 Chapter 16 Wetland Ecosystem Services

environment, such as the surface of a marsh, as well as a source of organic carbon,
something abundant in most wetlands. Because most temperate wetlands are the
receivers of fertilizer-enriched agricultural runoff and are ideal environments for
denitrification, it is likely that they are important to the world’s available nitrogen
balance. The phenomenon of nitrogen enrichment of coastal waters causing “dead
zones,” or hypoxia (dissolved oxygen <2.0mg/L in the hypolimnion now occurs
worldwide. (See “The Nitrogen Cycle, Wetlands, and Hypoxia” in Chapter 6:
“Wetland Biogeochemistry.”) Wetland restoration and creation in the watershed have
been recommended as a fundamental approach to solving this eutrophication (Mitsch
et ah, 2001; Mitsch and Day, 2006).

The global carbon cycle and wetlands are tightly linked. Wetlands, particularly
northern peatlands, have stored enormous quantities of carbon in the peat. When
these peatlands are protected and their water table is not affected, this carbon remains
essentially in storage forever. When this peat is oxidized, whether by burning direcdy
as a fuel or indirectly by altering the hydrology and causing drying and oxidation
of the peat, the peatlands could become important sources of carbon dioxide to the
atmosphere. Wetlands can be significant sinks of carbon if they are still building peat
or accumulating carbon in their soil. This could be a significant advantage for tropical
wetlands and for created and restored wedands that are sdll building carbon storage
in their soils compared to terrestrial systems that accumulate organic carbon in the
soil slowly. Wetland carbon sequestradon and greenhouse gas emissions are discussed
in more detail in Chapter 17: “Wetlands and Climate Change.”

Aquifer Recharge

Another value of wedands related to hydrology is groundwater recharge. This function
has received too litde attention, and the magnitude of the phenomenon has not been
well documented. Some hydrologists believe that, although some wedands recharge
groundwater systems, most wetlands do not. The reason for the absence of recharge is
that soils under most wedands are impermeable. In the few studies available, recharge
occurred primarily around the edges of wedands and was related to the edge : vol-
ume ratio of the wedand. Thus, recharge appears to be reladvely more important in
small wetlands, such as prairie potholes, than in large ones. These small wedands can
contribute significandy to recharge of regional groundwater.

Water Quality

Under favorable conditions, wetlands have been shown to remove organic and inor-
ganic nutrients and toxic materials from water that flows across them. The concept
of wetlands as sinks for chemicals was discussed in Chapter 6: “Wedand Biogeochem-
istry,” and the practice of using wedands for wastewater treatment and water quality
improvement is discussed in detail in Chapter 19: “Wedands and Water Quality.” Wet-
lands have six attributes that influence the chemicals that flow through them, whether
the chemicals are naturally added or ardficially applied:

Wetland Ecosystem Services 545

1 . Wetiands cause a reduction in water velocity as streams enter wetlands,
causing sediments and chemicals sorbed to sediments to drop out of the
water column;

2. Many anaerobic and aerobic processes occur in close proximity in wedands,
promoting denitrification, chemical precipitation, and other chemical
reactions that remove certain chemicals from the water;

3. High productivity in many wetlands can lead to high rates of mineral uptake
by vegetation and subsequent burial in sediments when the plants die;

4. A diversity of decomposers and decomposition processes occur in wetland
sediments;

5. There is a large contact surface of water with sediments because of the
shallow water, leading to significant sediment-water exchanges; and

6. Organic peat accumulates in many wetlands, causing the permanent burial of
chemicals.

Cultural Ecosystem Services

Aesthetics

A real but difficult aspect of a wetland to capture is its aesthetic value, often hidden
under the dry term nonconsumptive use values^ which simply means that people enjoy
being out in wetlands. There are many aspects of this kind of wetland use. Wetlands
are excellent “biological laboratories,” where students in elementary, secondary, and
higher education can learn natural history firsthand. They are visually and education-
ally rich environments because of their ecological diversity. Their complexity makes
them excellent sites for research. Many visitors to wetlands use hunting and fishing
as excuses to experience wildness and solitude, expressing that frontier pioneering
instinct that may lurk in all of us. In addition, wetlands are a rich source of informa-
tion about our cultural heritage. The remains of prehistoric Native American villages
and mounds of shells or middens have contributed to our understanding of Native
American cultures and of the history of the use of our wetlands.

Many artists — the Georgia poet Sidney Lanier, the painters John Constable and
John Singer Sargent, and others who paint and photograph wetlands — have been
drawn to them. Two artists — one a photographer and the other a painter — took a
one-year excursion through the wetlands of the Louisiana delta in 2004 and 2005
(Lockwood and Gary, 2005). Their works, shown as exquisite photographs and paint-
ings, have been shown in several museums throughout the United States.

Subsistence Use

In many regions of the world, the subsistence use of wetlands is extensive. There,
wetlands provide the primary resources on which village economies are based. These
societies have adapted to the local ecosystems over many generations and are inte-
grated into them. Some of these cultures, including the Camarguais in France, the

546 Chapter 16 Wetland Ecosystem Services

Louisiana Cajuns in the United States, and the Marsh Arabs in Iraq, are described in
Chapter 1: “Wetlands: Human Use and Science.”

Quantifying Ecosystem Services

Efforts have been made to quantify the “free services” and amenities that wedands
provide to society for more than 40 years. Starting with the economics via energy flux
approaches of H. T. Odum of the 1960s and 1970s that influenced a generation of
new scientists, publication of The Southern River Swamp — A Multiple-Use Environ-
ment (Wharton, 1970) and The Value of the Tidal Marsh (Gosselink et ah, 1974), a
significant literature now exists in the general field of ecological economics on ascrib-
ing values to wetlands for the services they provide. Costanza et al. (1997, 2014) took
these types of calculations one step further by estimating the public service functions
of all Earth’s ecosystems, including wetlands. These studies and others have gener-
ated a new vocabulary on ecosystem values with terms such as public service function,
natural capital, environmental services, and ecosystem jjoods and services. All of these
terms mean essentially the same thing. Nature, including wetlands, provides value to
humans, and the value needs to be recognized whenever wetlands are either threatened
or conserved (Sdderquist et ah, 2000; Mitsch and Gosselink, 2000).

Several approaches to the valuation of wetlands have been advanced. Because of
the complexities described previously, there is no universal agreement about which
approach is preferable. In part, the choice depends on the circumstances. Valuations
fall broadly into two classes: ecological (or functional) evaluation and economic (or
monetary) valuation. The former evaluation generally is necessary before attempting
the latter valuations; ecological functions are the causes of monetary values.

Ecological Valuation

Habitat Evaluation Procedures

Table 16.3 shows an example of the application of the Habitat Evaluation Proce-
dure (HEP) of the U.S. Fish and Wildlife Service to different development plans for
a cypress-gum swamp ecosystem. The present value of the swamp for a representa-
tive group of terrestrial and aquatic animals was evaluated (baseline condition) using a
habitat suitability index (HSI) based on a range of 0 to 1 for the optimum habitat for
the species in question. The evaluation resulted in a mean terrestrial HSI of 0.8 and
a mean aquatic HSI of 0.4. This baseline condition was compared with the projected
habitat condition in 50 and 100 years under three projected scenarios: Plan A, Plan
B, and a no-project projection. The results suggest that Plan A would be detrimental
to the environment, whereas Plan B would have no effect on terrestrial habitat val-
ues and would improve aquatic ones. Whether to proceed with either of these plans
is a decision that requires weighing the projected environmental effects against the
projected economic benefits of the project.

One often-neglected feature of the analysis is the effect of aggregating HSIs for
different species. Although, overall. Plan B appears to be about equivalent environ-
mentally to the no-project option, scrutiny of Table 16.3 shows that Plan B is expected
to improve the habitat for swamp rabbits and large-mouthed bass but decrease its value

Quantifying Ecosystem Services 547

Table 16.3 Habitat Evaluation Procedure of the impact of two management plans and a
no-management control in a cypress-gum swamp in southeastern USA^

Baseline

Condition

Future with
Project Plan tP

Future with
Project Plan B"

Future without
Project

Species

50 Years

100 Years

50 Years

100 Years

50 Years 100 Years

Terrestrial

Raccoon

0.7

0.5

0.6

0.8

0.7

0.9

Beaver

0.7

0.2

0.4

0.3

0.6

0.4

Swamp rabbit

0.7

0.2

0.8

0.7

0.4

Green heron

0.9

0.2

0.1

0.8

0.9

1.0

Mallard

0.8

0.3

0.2

1.0

0.9

1.0

Wood duck

0.8

0.3

0.2

0.9

1.0

Prothonotary warbler

0.8

0.3

0.1

0.6

0.7

0.8

0.9

Snapping turtle

0.8

0.4

0.3

0.8

0.7

0.8

0.9

Bullfrog

0.9

0.3

0.2

0.8

0.9

1.0

Total terrestrial HSI

7.1

2.7

2.1

6.9

7.0

7.4

7.5

Mean terrestrial HSI

0.8

0.3

0.2

0.8

Aquatic

Channel catfish

0.3

0.4

Largemouth bass

0.4

0.2

0.3

0.7

0.8

0.4

Total aquatic HSI

0.7

0.5

0.7

1.1

1.2

0.8

Mean aquatic HSI

0.4

0.3

0.4

0.6

0.4

^Numbers in the tables are habitat suitability index (HSi) values, which have a maximum value of 1 for an
optimal habitat.

'’Channelization of water and clearing of swamp for agricultural development with a loss of 324 ha of wetland.
"Construction of levees around swamp for flood control with no loss of wetland area.

Source: Schamberger et al. (1979)

for warblers and turtles. This kind of detailed scrutiny may be important because it
indicates a change in the quality of the environment, but it is often neglected when
the “apples and oranges” are combined into “fruit.”

Hydrogeomorphic Analysis

The hydrogeomorphic (HGM) classification described in Chapter 13: “Wetland Clas-
sification” also allows a quantification of the functions of wedands. Its uniqueness
lies in its quantification of natural wetland functions without regard to their signifi-
cance to society. This is done by comparing the wetland of interest to a reference site
that is characteristic of the same HCM class. Brinson et al. (1994) summarized the
assessment procedure:

1 . Group wetlands into HGM classes with shared properties. (The classification is
discussed in Chapter 13.)

2. Define the relationship between HGM properties and the functions of wetlands.
The goal is to select functions that are linked clearly and logically to wetland
HGM properties and that have hydrologic, geomorphic, and ecological
significance. This step represents the scientific basis for the presence of the
function.

548 Chapter 16 Wetland Ecosystem Services

3. Develop functional profiles for each wetland class. These can range from
descriptive narratives to multivariate data sets coveting numerous sites.

4. Develop a scale for expressinp; functions within each wetland class, by usinp;
indicators and profiles from the reference wetlands of that class. These scales
serve as benchmarks for each wetland class. Reference wetlands should
include the full range of natural and human-induced variations due to stress
and disturbance.

5. Develop the assessment methodology. The assessment relies on indicators to
reveal the likelihood that the functions being evaluated are present in the
wetland and depends on reference populations to scale the assessment. The
reference wetlands are also used to set goals for compensatory mitigation.

Evaluating Alternatives with the HGM Technique — Illustration from

North Carolina

In an illustration of the method to estimate the impact of a project or restore-

tion on wetland functions,

Rhelnhardt et al. (1997) apply the HGM method

to evaluate mitigation strategies in mineral soil forested pine (Pinus palustris)

Table 16.4 Field parameters used to estimate ecosystem function in a

hydrogeomorphic assessment of forested wetlands in southeastern

North Caroiina

Variable

Description

Hydrology/Topography

^DITC

Lack of ditches nearby (<50 m)

'^MICR

Microtopographic complexity

Herbaceous Vegetation

^GRAM

Percentage cover of graminoids

^FORB

Percentage cover of forbs

Canopy Vegetation

'^REE

Total basal area for trees (m^/ha; >10 cm DBH)

'hoEN

Density of canopy trees (stems/ha; >10 cm DBH)

^TDIA

Average tree diameter (m)

^CVEG

Sorensen simiiarity index of canopy importance value

Subcanopy Vegetation

'^SUBC

Density of subcanopy (stems/ha)

'^SDLG

Percentage cover of trees and shrubs <1 m tall

^SVEG

Sorensen similarity index of subcanopy importance value

Litter/Standing Dead

'^LTR

Litter depth (cm)

'^SNAG

Density of standing dead stems (stems/ha)

^CWD

Volume of coarse woody debris (cm^/ha)

Source: Rhelnhardt et al. (1997)

Quantifying Ecosystem Services 549

Table 16.5 Predicted changes in hydrologic regime function resuiting from a
hypotheticai airport construction on one wetland site and the comparison of the
mitigation required for two different wetland restoration alternatives (variabies
are defined in Table 16.4)

Wetland

Being

Restoration

Destroyed

Alternataive

Alternative 2'^

Reference

After

Wetland

Now Airport

Now Restoration

Variable

Raw

Index

Raw

Index Raw

Index

Raw Index

Raw

Index

Raw

Index

Raw

Index

''tree

14.7

1.0

14.6

1.0 —

0.0

15.3

1.0

10.0

0.7

12,550

1.0

13,314

1.0 —

0.0

6,963

0.5

18,402

0.5

9,800

0.8

2.5

1.0

2.5

1.0 —

0.0

1.0

4.2

1.0

4.2

1.0

0.5

0.5 —

0.0

1.0

0.5

1.0

Functional

1.0

0.71

0.0

0.71

0.64

0.91

index'^

Relative -0.71 +0.71 +0.27

impact

Mitigation 0.71/0.71 = 1:1 0.71/0.27 = 2.6:1

ratio"

"Restoration of an agricultural field (former wetland) to a forested wetland.

'’Restoration of a pine plantation to a forested wetland.

‘’Hydrologic functional index = +V'subc +^micr)/3> ±

‘'Ratio of wetland must be restored to area of wetland destroyed to achieve functional equivalent
hydrologic regime.

Source: Rheinhardt et al. (1997).

flats in North Carolina. Fourteen variables were used to estimate the function
of both study and reference wetlands (Table 16.4). Absolute values of some
of the variables (e.g., tree density) are then translated into indices on a scale
of 0.0 to 1.0 by comparing those functions to a reference wetland site. Such
indices, in turn, are applied to model functions, such as “maintain hydrologic
regime” as in Table 16.5, and comparisons of human impact on wetlands can
be assessed. Table 16.5 shows a hypothetical case in which an airport is
destroying a wetland (with an overall loss index of 0.71 when compared to a
nearby reference, which, by definition, has an index of 1.0), and two restora-
tion alternatives are being considered. The analysis shows that restoration of
a cropland back to a wetland (Restoration Alternative 1) would be a good alter-
native because the cropland currently has 0.0 value in maintaining hydrologic
regime. Thus, the restoration is estimated to require only 1 ha of that crop-
land (gain = -1-0.71) for every hectare of wetland lost due to the airport (loss
= -0.71). This is a 1:1 mitigation ratio (the ratio of area of wetland restored
to wetland lost).

550 Chapter 16 Wetland Ecosystem Services

Restoration of an existing pine plantation to a natural pine wetland, how-
ever, would probably be easier but, functionally, the plantation already has
some of the desired values of wetlands. (It rates a functional index of 0.64
before any restoration takes place and would rate an index of 0.91 after
restoration, a net change of -1-0.27.) So that restoration strategy would require
2.6 ha (0.71/0.27) of pine plantation to be restored for every hectare of wet-
land lost for the airport (mitigation ratio = 2.6:1).

Economic Evaluation

Evaluation systems that seek to compare natural wetlands to human economic sys-
tems usually reduce all values to monetary terms (thus losing sight of the apples and
oranges). Conventional economic theory assumes that in a free economy, the eco-
nomic benefit of a commodity is the dollar amount that the public is willing to pay
for the good or service rather than be without it.

Although this characterization of value is reasonable under most conventional
economic conditions, it leads to real problems in monetizing nonmarket commodi-
ties, such as pure water and air, and in pricing wetlands whose value in the marketplace
is determined by their value as real estate, not by their “free services” to society. Con-
sequently, attempts to monetize wetland values have generally emphasized the com-
mercial crops from wedands: fish, shellfish, furs, and recreational fishing and hunting,
for which pricing methodologies are available. This kind of pricing ignores ecosystem-
and global-level ecosystem services related to clean air and water and other life-support
functions. Even in the cases of market commodities from wetlands, available data are
seldom adequate to develop reliable demand curves.

Economists recognize four more or less independent aspects of “value” that con-
tribute to the total. These aspects are:

1 . Use value. The most tangible portion of total value derived from identifiable
direct benefits to the individual; hunting, harvesting fish, and nature study
are examples.

2. Social value. Those amenities that accrue to a societal group rather than an
individual; examples are improved water quality, flood protection, and the
maintenance of the global sulfur balance.

3 . Option value. The value that exists for the conservation of perceived benefits
for future use.

4. Existence value. The benefits deriving from the simple knowledge that the
valued resource exists, irrespective of whether it is ever used. For example,
the capacity of an extant wetland to conserve biological diversity is an
existence value.

Quantifying Ecosystem Services 551

As we have seen, use value is the easiest to estimate. The other three values, which are
more difficult to quantify and also generally reflect longer-term viewpoints, have been
addressed by economists using alternative methods, as illustrated next.

Willingness-to-Pay Methods

In the absence of a well-developed free-market alternative, pricing methodologies have
been applied. One of these, wilUn£iness to pay, establishes a more or less hypothetical
(contingency) market for nonmarket goods or services. Willingness to pay or, more
accurately, net willingness to pay, is “the amount society would be willing to pay to
produce and/or use a good beyond that which it actually does pay” (Scodari, 1990).
The principle is illustrated as follows: Suppose a fisherman were willing to pay $30 a
day to use a particular fishing site but had to spend only $20 per day in travel and
associated costs. The net benefit, or economic value, to the fisherman of a fishing day
at the site is not the $20 expenditure but the $10 difference between what he was
willing to spend and what he had to spend. If the fishing opportunity at the site was
eliminated, the fisherman would lose $10 worth of satisfaction fishing; the $20 cost
that he would have incurred would be available to spend elsewhere. In the case of
commercial goods, such as harvested fish, the total value of a wetland is the sum of
the net benefit to the consumer plus the net benefit to the producer (the fisherman).

Opportunity Costs

A second approach to resource evaluation in the absence of a free-market model is
the opportunity cost approach. In general terms, the opportunity cost associated with
a resource is the net worth of that resource in its best alternative use. For example,
“the opportunity cost of conserving a wedand area is the net benefit which might have
been derived from the best alternate use of the area which must be foregone in order
to preserve it in its natural state” (Bardecki, 1987). Because determining the oppor-
tunity cost associated with wetland conservation would require the evaluation of each
wetland service as well as the identification and valuation of the best alternative use, in
practice, a comprehensive evaluation of the opportunity cost of wetland conservation
is far from possible. Nevertheless, it may represent a useful approach to the valuation
of specific wetland functions.

Replacement Value

If one could calculate the cheapest way of replacing various services performed by a
wetland and could make the case that those services would have to be replaced if the
wetland was destroyed, then the figure arrived at would be the replacement value.
Some of the replacement technologies that might be necessary to replace services
provided by wetland processes are listed in Table 16.6. A sample calculation of the
replacement cost method is shown in Table 16.7. In this example, a fish hatchery is
used to calculate fishery production, a flood reservoir to calculate flood and drought
control, sediment dredging to estimate sediment retention, and wastewater treatment
to estimate water quality enhancement.

552 Chapter 16 Wetland Ecosystem Services

Table 16.6 Some replacement technologies for societal support values provided by wetlands

Societal Support

Replacement Technologies

Peat Accumulation

Accumulating and storing organic matter (peat)

Hydrologic Functions

Maintaining drinking water quality

Maintaining groundwater level
Maintaining surface water level

Moderation of water flows

Biogeochemicai Functions

Processing sewage; cleansing nutrients and chemicals

Maintaining drinking water quality

Filter to coastal waters

Food Chain Functions

Providing food for humans and domestic animals
Providing cover

Sustaining anadromous trout populations

Sustaining other fish species and wetland-dependent
flora and fauna

Species diversity; storehouse for genetic material
Bird watching, sport fishing, boating, and other
recreational values
Aesthetic and spiritual values

Artificial fertilizers
Artificial flooding

Water transport

Pipeline to distant source

Well-drilling

Saltwater filtering

Dams for irrigation

Pumping water to dam

Irrigation pipes and machines

Water transport for domestic animals

Regulating gate

Pumping water to stream

Mechanical sewage treatment

Sewage transport

Sewage treatment plant

Clear-cutting ditches and stream

Water quality inspections

Water purification plant

Silos for manure from domestic animals

Nitrogen filtering

Water transport

Nitrogen reduction in sewage treatment plants

Agriculture production

Import of food

Roofing materials

Releases of hatchery-raised trout

Farmed salmon

Work by nonprofit organizations

Replacement not possible
Replacement not possible

Replacement not possible

Source: Folke (1991)

This approach has the merit of being accepted by some conventional economists.
For certain functions, it gives very high values compared with those of other valuation
approaches discussed in this section. For example, the tertiary treatment of wastewater
is extremely expensive, as is the cost of replacing the nursery function of marshes for
juvenile fish and shellfish. Serious questions, however, have been raised about whether
these functions would be replaced by treatment plants and fish nurseries if the wedands

Quantifying Ecosystem Services 553

Table 16.7 Estimated value of 770-ha riparian wetlands along the Kankakee River,
northeastern Illinois, estimated by repiacement value approach and by energy analysis

Replacement Cost Approach

$/Year

Total Value

Ecosystem Function (Replacement Technology)

Fish productivity (fish hatchery)

$91,000

Flood control/drought prevention (flood control reservoir)

$691,000

Sediment control (sediment dredging)

$100,000

Water quality enhancement (wastewater treatment)

$57,000

Total replacement cost

$939,000

Value/area $939,000 yrV770 ha =

U.S. $1,219 ha-i yr-i

Energy Flow Approach

Energy Flow Parameter

Number

Total Value

Ecosystem gross primary productivity (kcal m“^ yr“^)

20,000

Energy quality conversion, (kcal GPP/kcal fossil fuel)

Energy conversion in U.S. economy (kcal fossil fuel /U.S.$)

14,000

Value/area =

U.S. $714 ha-i yr-i

kcai = kilocalorie; GPP = gross primary productivity.
Source: Mitsch et al. (1979)

were destroyed. Some ecologists and economists argue that, in the long run, either
the services of wetlands would have to be replaced or the quality of human life would
deteriorate. Other individuals argue that this assertion cannot be supported in any
convincing manner.

Energy Analysis

A completely different approach uses the idea of energy flow through an ecosystem or
the similar concept of embodied energy. The concepts of embodied energy (Costanza,
1980), and emer^y (= energy memory; H. T. Odum, 1988, 1989, 1996) both attempt
to estimate the total energy required to produce something and then translate the
energy analysis into economic terms. It is assumed to be a valid index of the totality
of ecosystem functions and is applicable to human systems as well. In this way, both
natural and human systems can be evaluated on the basis of one common currency:
energy. Because there is a clear relationship between energy and money in our society,
energy flow can be translated to the more familiar currency of dollars at the end of the
evaluation.

A simple calculation using the annual energy flow of a bottomland forested wet-
land in Illinois is illustrated in Table 16.7. Here, an estimated ecosystem energy flow
(gross primary productivity [GPP]) of 20,000 kcal m^^yr^'^ yielded an estimated value
of $714 ha“^yr“^ . The energy analysis method gave a number about 60 percent of the
replacement value. The concept of energy “quality” was used in this calculation to
differentiate between energy flow in the ecosystem (based on gross primary produc-
tivity) and energy flow in the human-based fossil fuel economy. This is a precursor to
the current approach of emergy discussed below.

554 Chapter 16 Wetland Ecosystem Services

Louisiana Coastai Wetiands: Comparing Energy and Economic
Anaiyses

Costanza et al. (1989) showed that the economist’s willingness-to-pay
approach and energy analysis converge to a surprising degree for coastal
marshes in Louisiana, although both methods result in a great deal of
uncertainty (Table 16.8). The energy analysis approach yielded higher wetland
values, but the ranges overlap. The sensitivity of both conventional and energy
analysis methods to the choice of a discount rate, which has been vital for
decades in the outcome of cost-benefit studies, is also demonstrated in this
comparison. The energy analysis method is based on using the total amount
of energy captured by natural ecosystems as a measure of their ability to do
useful work (for nature and hence for society). The gross primary productivity
(GPP) of representative coastal marsh systems, which ranges from 48,000
to 70,000 kcal m“^ yr^, is converted to monetary units by multiplying by a
conversion factor of 0.05 units fossil fuel energy/unit GPP energy and dividing
by the energy/money ratio for the economy (15,000 kcal fossil fuel/1983 $).
These calculations resulted in an estimate of annual coastal wetland value
of about $1,560 ha~^ yr^, which, when converted to present value for an
infinite series of payments, yields the range of capitalized values of $16,000
to $70,000 ha“^ for the discount rates used in Table 16.8.

Table 16.8 Estimates of wetland values in $/ha of Louisiana coastal
marshes based on willingness-to-pay and energy anaiysis at two
discount rates

Discount Rate

Method

Willingness to pay
Commercial fishery

$2,090

$783

Fur trapping (muskrat and nutria)

991

373

Recreation

447

114

storm protection

18,653

4,732

Total willingness-to-pay value

$22,181

$6,002

Energy analysis

$42,000-$70,000

$16,000-$26,000

Best estimate

$22,000-$42,000

$6,000-$16,000

Source: Costanza et al. (1989)

In comparison, the willingness-to-pay estimates reflect the assessment
that a reasonable range of wetland value for coastal Louisiana is between
$6,000 and $22,000 ha~^, depending on the discount rate applied to

Quantifying Ecosystem Services 555

determine the present value. Costanza et al. (1989) used this range from
the willingness-to-pay and energy analysis approaches to suggest that the
annual loss of Louisiana coastal wetlands is costing society from $77 million
to $544 million per year.

Emergy Analysis

Emcrgy analysis is a variation on the energy analysis (both terms were pioneered by
H. T. Odum at the University of Florida in the 1970s and 1980s). The key to emergy
analysis is the determination of transformities, or ratios that allow the conversion of
one form of energy to another, as was done previously for gross primary produc-
tivity and fossil fuel energy described in the example above. These ratios are usually
expressed in terms of solar emjoules (sej) per joule (or similar unit) of base energy or
ecosystem flow. An example of an emergy flow analysis used for wedands is illustrated
in the next box.

Emergy Analysis of Wetlands in Florida

A comparison was made among three types of wetlands in Florida — a forested
wetland, a shrub-scrub wetland, and a marsh (Bardi and Brown, 2001) — to
compare their ecosystem services. The services considered were not only
gross primary productivity but also infiltration of water to the groundwater
(groundwater recharge) and transpiration. In addition, the storages of natu-
ral capital (stored water, biomass, and basin structure) were added. When all
of the environmental services and natural capital are first converted to solar
emjoules (sej) and then to dollars (Table 16.9), the data suggest that a 1-ha

Table 16.9 Results of emergy analysis comparing the economic
value of three types of wetlands in Florida for their environmentai
services and natural capital (Values are U.S.$/ha)

Ecosystem Type

Environmental

Services’^

Natural Capital*’

Total Value

Forested wetland

$231,880

$1,322,723

$1,554,603

Shrub/Scrub wetland

$31,831

$1,075,536

$1,107,366

Freshwater marsh

$13,173

$626,645

$639,817

®Environmental services include gross primary productivity, infiltration, and
transpiration.

‘’Natural capital includes live biomass, peat, water, and basin structure (formed
by geological processes).

Source: Bardi and Brown (2001)

556 Chapter 16 Wetland Ecosystem Services

forested wetland is approximately 2.4 times more valuable than a similar-size
marsh. Furthermore, the analysis points out that the wetland values range
from $640,000 to $1.5 million per ha. At the time, the going rate for buying
wetland mitigation credit in Florida was $187, 000/ha. Thus, the rate being
paid for mitigation credit was one-third to one-eighth that of the values calcu-
lated for these wetlands. According to this estimate, wetlands were being sold
to destruction at too low a price.

Energy and emergy analyses, although imprecise because of the many con-
version factors needed, are more satisfying to many scientists than conventional
cost- accounting methods, because they are based on the inherent function of the
ecosystem, not on perceived values that may change from generation to generation
and from location to location.

Valuing Ecosystem Goods and Services

Costanza et al. ( 1997) wrote a highly cited paper on the value of the goods and services
from ecosystems and suggested that the world’s ecosystems were worth $33 trillion per
year (1995 $). Updated, this is equivalent to U.S.$46 trillion per year (in 2007 U.S.$).
That study used ecosystem unit estimators that showed that wetlands, especially inland
swamps and floodplains, were considerably more valuable than lakes and rivers, forests,
and grasslands ( Table 1 6 . 1 0 ) . Only coastal estuaries had higher unit values than inland
and coastal wetlands from the 1997 study.

Balmford et al. (2002) argued that the mt marginal benefits of ecosystems should
be estimated rather than the aggregated numbers developed by Costanza et al. ( 1997),
which often were simple replacement values. The net marginal benefit is the differ-
ence between values of relatively intact ecosystems and the values to humans of the
same ecosystems converted to human use. After investigating over 300 case studies.

Table 16.10 Estimated unit vaiues of ecosystems (aii numbers
nermalized te 2007 U.S.$)

Ecosystem

1997 Estimate
Unit Value
(U.S.$ ha-i yr-i)

2011 Estimate
Unit Value
(U.S.$ ha-i yr-i)

Estuaries

31,509

28,916

Inland swamps/floodplalns

27,021

25,681

Tidal marshes/mangroves

13,786

193,843

Lakes/rIvers

11,727

12,512

Forests

1,338

3,800

Grasslands

321

4,166

Source: Costanza et al. (2014); 1997 estimates from Costanza et al. (1997) but
revised to 2007U.S.$

Quantifying Ecosystem Services 557

Balmford et al. (2002) came up with only five studies worldwide where economic
estimates were available for both conditions — intact ecosystems and the same land-
scape heavily managed. Two of those five case studies were of wetlands (Fig. 16.10).
An economic analysis of a mangrove swamp in Thailand showed that conversion of a
swamp to aquaculture made economic sense in the short term, but in the long term,
the total economic value of an intact mangrove swamp was $60,400, about 3.6 times
that of the value of converting the swamp to shrimp aquaculture. The values provided
by the natural mangrove swamp included timber, charcoal, nontimber forest products,
offshore fisheries, and storm protection. In a similar comparison, a freshwater marsh
in Canada was found to have a total economic value of $8, 800/ha, about 2.4 times
the value realized by converting the wetland to intensive agriculture. Here, the major
values of the natural marsh were for sustainable hunting, fishing, and trapping. The
Balmford et al. (2002) estimates were used extensively by the Millennium Ecosystem

80,000

>•

60,000

40,000

20,000

Mangrove swamp, Thailand

6 =

6% over

30 years 1
1
1
1
1
1
1
1

Intact swamp Shrimp farming

10,000

^ 8,000

^ 6,000

g;' 4,000
z

2,000

Intact wetland Intensive farming

Wetland, Canada
6 = 4% over 50 years

Figure 16.10 Two case studies of the marginai benefits of natural wetlands versus con-
version of the wetland to intensive human industry: (a) mangrove system in Surat Thani,
southern Thaiiand, and (b) freshwater marshes in Canada, d indicates discount rates; NPV
indicates net present vaiue in year 2000 U.S.$/ha. (From Balmford et al., 2002)

558 Chapter 16 Wetland Ecosystem Services

Assessment (2005) that was published a few years later, two of only four case stud-
ies worldwide featured to show that sustainably managed ecosystems provide more
economic benefit than do conversion to agricultural and aquaculture uses.

Costanza et al. (2014) revisited the calculations of his 1997 paper, using some
revised unit values determined by deGroot et al. (2012) and others from a new United
Nations-sponsored post-Millennium Ecosystem Assessment. The unit values for
ecosystems are given in the last column in Table 16.10. The inland swamps/floodplain
number stayed approximately the same while the tidal marsh/mangroves unit value
increased 14-fold, “largely due to new studies of the storm protection, erosion
protection, and waste treatment values” of these tidal wetlands. The overall value of
ecosystems of the world, updated from the Costanza et al. (1997) paper, is U.S.$125
trillion to $145 trillion per year (Costanza et al., 2014). This reinvestigation of “what
nature is worth” has received significant press coverage (e.g., Zimmer, 2014) in the
New York Times^ and Rosen (2014) in The Atlantic?

Problems and Paradoxes of Quantifying Wetiand Vaiues

Regardless of which kind of ecosystem evaluation is used, eight generic problems and
paradoxes to quantifying wetland values should be appreciated:

1. The terms value and service are anthropocentric; hence, assigning values to
different natural processes usually reflects human perceptions and needs
rather than intrinsic ecological processes.

2. The most valuable products of wetiands are public amenities that have no
commercial value for the private wetland owner.

3. The ecological value, but not necessarily the economic value, of a wetland
depends on its context in the landscape.

4. The relationships among wedand area, surrounding human population, and
marginal value are complex.

5. Commercial values are finite, whereas wetlands provide values in perpetuity.

6. A comparison of economic short-term gains with wedand value in the long
term is often not appropriate.

7. Estimates of values and services, by their nature, are colored by the biases of
individuals and society and by the economic system.

8 . A landscape view of wetlands is required to make intelligent decisions about
the values of created and managed wedands.

If one ignores the technical problems of funcdonal ecosystem substitudon, the
idea attracts many people because of the common perception among economists that
any commodity can be replaced. As scarcity of one product drives the price up, the

'www.nytimes.com/2014/06/05/science/earth/putting-a-price-tag-on-natures-defenses.html?_r=2.

2www.theatlantic.eom/business/archive/2014/06/how-much-are-the-worlds-ecosystems-worth/

372862/.

References 559

creativity of the free market will surely result in the development of a cheaper sub-
stitute. This is not true of ecosystems, however. Much of the value of an ecosystem,
especially an open system such as a wetland, depends on its landscape context and on
strong interactions among the parts of the landscape. Thus, the value of a riparian for-
est depends on its ecological links to the adjacent stream on one side and the upland
fields or forest on the other.

Faustian Bargain

Because of the many problems documented in this chapter relate to valuation of nat-
ural ecosystem services, many ecologists oppose economic valuation of ecosystems. It
implies that natural systems can be equated in the marketplace to other market prod-
ucts. Attempts to place dollar values on natural ecosystems, however, such as those
cited in this chapter, have raised public awareness (see, e.g., Zimmer, 2014; Rosen,
2014) of the high value of the goods and services of nature, and in this way helped in
efforts to conserve natural resources. Thus, ecologists are caught in a Faustian bargain
with the devil, trying to make the case for ecosystem services in the common currency
of our civilization while clearly documenting the reasons why natural ecosystem con-
servation should not depend on the operation of free-market forces. There is no easy
answer to this dilemma.

Recommended Readings

Costanza, R., R. de Groot, R Sutton, S. van der Ploeg, S. J. Anderson, I. Kubiszewski,
S. Farber, and R. K. Turner. 2014. Changes in the Global Value of Ecosystem
Services. Global Environmental Chan£ie 26\ 152-158.

Millennium Ecosystem Assessment. 2005. Ecosystems and Human Well-Being: Synthe-
sis. Washington, DC: Island Press.

References

Balmford, A., A. Bruner, P. Cooper, R. Costanza, S. Farber, R. E. Green, M. Jenkins,
P. Jefferiss, V. Jessamy, J. Madden, K. Munro, N. Myers, S. Naeem, J. Paavola,
M. Rayment, S. Rosendo, J. Roughgarten, K. Trumper, and R. K. Turner. 2002.
Economic reasons for conserving wild nature. Science 297: 950-953.

Barbier, E. B., I. Y. Georglou, B. Enchelmeyer, and D. J. Reed. 2013. The value of
wetlands in protecting Southeast Louisiana from hurricane storm surges. PLoS
ONES (3): e58715.

Bardecki, M. J. 1987. Wetland Evaluation: Methodology Development and Pilot Area
Selection. Report 1, Canadian Wildlife Service and Wildlife Habitat Canada,
Toronto.

Bardi, E., and M. T. Brown, 2001. Emergy evaluation of ecosystems: A basis for
environmental decision making. In M. T. Brown, ed., Emergy Synthesis: Theory
and Applications of the Emergy Methodology. Proceedings of a conference held at
Gainesville, FL, September 1999. Center for Environmental Policy, University of
Florida, Gainesville, pp. 81-98.

560 Chapter 16 Wetland Ecosystem Services

Brinson, M. M., W. Kruczynsld, L. C. Lee, W. L. Nutter, R. D. Smith, and D. F.
Whigham. 1994. Developing an approach for assessing the functions of wetlands.
In W. J. Mitsch, ed.. Global Wetlands: Old World and New. Elsevier, Amsterdam,
the Netherlands, pp. 615-624.

Cahoon, D. R., D. J. Reed, and J. W. Day, Jr. 1995. Estimating shallow subsidence in
microtidal salt marshes of the southeastern United States — Kaye and Barghoorn
revisted. Marine Geology 128: 1-9.

Check, E. 2005. Roots of recovery. Nature 438: 910-911.

Costanza, R. 1980. Embodied energy and economic evaluation. Science 210:
1219-1224.

Costanza, R., S. C. Farber, and J. Maxwell. 1989. Valuation and management of wet-
land ecosystems. Ecological Economics 1: 335-361.

Costanza, R., R. d’Arge, R. de Groot, S. Farber, M. Grasso, B. Elannon, K. Limburg,
S. Naeem, R. V. O’Neill, J. Paruelo, R. G. Raskin, P. Sutton, and M. van den Belt.
1997. The value of the world’s ecosystem services and natural capital. Nature 387 :
253-260.

Gostanza, R., W. J. Mitsch, and J. W. Day. 2006. A new vision for New Orleans and
the Mississippi delta: Applying ecological economics and ecological engineering.
Erontiers in Ecolojjy and the Environment 4: 465-472.

Gostanza, R., R. de Groot, P. Sutton, S. van der Ploeg, S. J. Anderson, I. Kubiszewski,
S. Farber, and R. K. Turner. 2014. Ghanges in the global value of ecosystem
services. Global Environmental Change 26\ 152-158.

Danielsen, F., M. K. Sorensen, M. F. Olwig, V. Selvam, F. Parish, N. D. Burgess, T.
Hiraishi, V. M. Karunagaran, M. S. Rasmussen, L. B. Hansen, A. Quarto, and N.
Suryadiputra. 2005. The Asian tsunami: A protective role for coastal vegetation.
Science 310: 643.

Das, S, and A-S. Grepin. 2013. Mangroves can provide protection against wind during
storms. Estuarine, Coastal and Shelf Science 134: 98-107.

Day, J. W., Jr., J. Barras, E. Glairain, J. Johnston, D. Justic, G. P. Kemp, J.-Y. Ko, R.
Lane, W. J. Mitsch, G. Steyer, P. Templet, and A. Yahez-Arancibia. 2005. Implica-
tions of global climatic change and energy cost and availability for the restoration
of the Mississippi Delta. Ecolo£iical Engineering 24: 253-265.
deGroot, R., Brander, L., van der Ploeg, S., Gostanza, R., Bernard, F., Braat, L.,
Ghristie, M., Grossman, N., Ghermandi, A., Hein, L., Hussain, S., Kumar, P,
McVittie, A., Portela, R., Rodriguez, L.G., ten Brink, P, and van Beukering, P.
2012. Global estimates of the value of ecosystems and their services in monetary
units. Ecosystem Services 1: 50-61.

Feierabend, S. J., and J. M. Zelazny. 1987. Status report on our nation’s wedands.

National Wildlife Federation, Washington, DG. 50 pp.

Folke, G. 1991. The societal value of wedand life-support. In G. Folke and T.
Kaberger, eds.. Linking the Natural Environment and the Economy. Kluwer
Academic Publishers, Dordrecht, the Netherlands, pp. 141-171.

Food and Agriculture Organizadon of the United Nations (FAO). 2003. State of the
World’s Forests. FAO, Rome, Italy.

References 561

Gibbons, A. 1997. Did birds fly through the K-T extinction with flying colors.^ Science
275: 1068.

Gosselink, J. G., E. P. Odum, and R. M. Pope. 1974. The value of the tidal marsh.
Publication LSU-SG-74-03. Genter for Wetland Resources, Louisiana State Uni-
versity, Baton Rouge. 30 pp.

Harris, L. D., and J. G. Gosselink. 1990. Gumulative impacts of bottomland hard-
wood forest conversion on hydrology, water quality, and terrestrial wildlife. In J.
G. Gosselink, L. G. Lee, and T. A. Muir, eds.. Ecological Processes and Cumula-
tive Impacts. Illustrated by Bottomland Hardwood Wetland Ecosystems. Lewis
Publishers, Ghelsea, Ml, pp. 259-322.

Harter, S. K., and W. J. Mitsch. 2003. Patterns of short-term sedimentation in a fresh-
water created marsh. Journal of Environmental Quality 32: 325-334.

Hiraishi, T., and K. Harada. 2003. Greenbelt tsunami prevention in south-Paciflc
region. Report of the Port and Airport Research Institute, vol. 24, no. 2.
EqTAP Project, Japan, www.drs.dpri.kyoto-u.ac.jp/eqtap/report/papua/
tsunam_prevention / tsunami_pari_report3 .pdf

Lockwood, G. G., and R. Gary. 2005. Marsh Mission: Capturing the Vanishinp; Wet-
lands. Louisiana State University Press, Baton Rouge. 106 pp.

Louisiana Goastal Wedands Gonservation and Restoration Task Force and the
Wetlands Gonservation and Restoration Authority. 1998. Coast 2050: Toward a
Sustainable Coastal Louisiana, an Executive Summary. Louisiana Department of
Natural Resources, Baton Rouge. 12 pp.

Marois, D. and W. J. Mitsch. 2015. Goastal protection from tsunamis and cyclones
provided by mangrove wetlands — a review. International Journal for Biodiversity
Science, Ecosystems Services and Manap/ement, doi.org/10. 1080/21 5 13732. 2014
.997292

Millennium Ecosystem Assessment, 2005. Ecosystems and Human WelTBein^: Synthe-
sis. Island Press and World Resources Institute, Washington, DG. 137 pp.

Mitsch, W. J., M. D. Hutchison, and G. A. Paulson. 1979. The Momence Wetlands of
the Kankakee River in Illinois — an assessment of their value. Document 79/17,
Illinois Institute of Natural Resources, Ghicago. 55 pp.

Mitsch, W. J., and J. G. Gosselink. 2000. The value of wetlands: Importance of scale
and landscape setting. Ecological Economics 35: 25-33.

Mitsch, W. J., and J. W. Day, Jr. 2006. Restoration of wetlands in the Mississippi-Ohio-
Missouri (MOM) River Basin: Experience and needed research. Ecological Enpfi-
neerinp! 26: 55-69.

Niering, W. A. 1988. Endangered, threatened and rare wetland plants and animals
of the continental United States. In D. D. Hook et al., eds.. The Ecology and
Management of Wetlands, Vol. 1: Ecolopty of Wetlands. Timber Press, Portland, OR,
pp. 227-238.

562 Chapter 16 Wetland Ecosystem Services

Novitzki, R. R 1985. The effects of lakes and wetlands on flood flows and base flows in
selected northern and eastern states. In H. A. Groman, D. Burke, T. Henderson,
and H. Groman, eds.. Proceedings of a Conference — Wetlands of the Chesapeake.
Environmental Law Institute, Washington, DG, pp. 143-154.

Odum, H. T. 1988. Self-organization, transfbrmity, and information. Science 242:
1132-1139.

Odum, H. T. 1989. Ecological engineering and self-organization. In W. J. Mitsch and
S. E. Jorgensen, eds., Ecolopfical Engineering . John Wiley 8c Sons, New York, pp.
79-101.

Odum, H. T. 1996. Environmental Accounting: Energy and Environmental Decision
Making. John Wiley 8c Sons, New York. 370 pp.

Ogawa, H., and J. W. Male. 1983. The Flood Mitipiation Potential of Inland Wetlands.
Publication 138, Water Resources Research Genter, University of Massachusetts,
Amherst. 164 pp.

Ogawa, H., and J. W. Male. 1986. Simulating the flood mitigation role of wetlands.
Journal of Water Resource Planning and Manap/ement 112: 114-128.

Rheinhardt, R. D., M. M. Brinson, and P. M. Farley. 1997. Applying wetiand reference
data to functional assessment, mitigation, and restoration. Wetlands 17: 195-215.

Rosen, R. J. 2014. How much are the world’s ecosystems worth.> The Atlantic., June
16, 2 pp. www.theatlantic.eom/business/archive/2014/06/how-much-are-the-
worlds-ecosystems-worth/372862/.

Schamberger, M. L., G. Short, and A. Farmer. 1979. Evaluation wedands as a wildlife
habitat. In P. E. Greeson, J. R. Glark, and J. E. Glark, eds.. Wetland Functions and
Values: The State of Our Understanding . American Water Resources Association,
Minneapolis, MN, pp. 74-83.

Scodari, P F. 1990. Wetlands protection: The role of economics. Environmental Law
Institute, Washington, DG. 89 pp.

Sdderqvist, T., W. J. Mitsch, and R. K. Turner, eds. 2000. The Values of Wetlands:
Landscape and Institutional Perspectives. Special issue of Ecological Economics 35:
1-132.

Turner, R. E. 1982. Protein yields from wetlands. In B. Gopal, R. E. Turner, R. G.
Wetzel, and D. F. Whigham, eds.. Wetlands: Ecology and Manapfement . National
Institute of Ecology and International Scientific Publications, Jaipur, India,
pp. 405-415.

U.S. Army Gorps of Engineers. 1972. Gharles River Watershed, Massachusetts. New
England Division, Waltham, MA. 65 pp.

Weller, M. W. 1999. Wetland Birds. Gambridge University Press, Gambridge, UK.

Wharton, G. H. 1970. The Southern River swamp — A multiple-use environment.
Bureau of Business and Economic Research, Georgia State University, Atlanta.
48 pp.

Zimmer, G. 2014, June 5. Putting a price tag on nature’s defenses. New York Times.
http:/ /www.nytimes.com/2014/06/05/science/earth/putting-a-price-tag-
on-natures-defenses.html.Ur=l .

Chapter IT

Wetlands and Climate Change

Earth’s climate is chan£iin£i, as witnessed by higher atmospheric temperatures,
decreased snow and ice cover, and increasinpi sea levels, especially for the last 30
or 40 years. Wetlands emit 20 to 25 percent of global methane emissions to
Earth’s atmosphere, yet they also have the best capacity of any ecosystem to retain
carbon through permanent burial (sequestration). Both processes have
implications for climate change. Of the total storage of organic carbon in
Earth’s soils, 20 to 30 percent or more is stored in wetlands, and that storage is
vulnerable to loss back to the atmosphere if the climate warms or becomes drier.

Our estimates show that the world’s wetlands are climate-change positive; that
is, the negative effects of methane emissions on climate are more than
compensated for by carbon sequestration into peat or wetland soils.

The effects of climate change on coastal wetlands could be significant if sea
level rises, particularly in large river deltas where land subsidence is already
occurring and where inland migration of wetlands is prevented by human
development in a process called coastal squeeze. Eor inland wetlands, change in
precipitation patterns and warmer temperatures likewise can have detrimental
effects on wetland function.

Wetlands have significant yet still underappreciated roles in the global carbon cycle.
They are also positioned in the landscape where climate change could affect them
more than most other ecosystems. So their roles both as players in and recipients of
climate change are the subject of this chapter.

Climate Change

There is litde doubt that something significant is happening to our climate. Accord-
ing to the consensus of hundreds of scientists who have been involved in the

563

564 Chapter 17 Wetlands and Climate Change

Year

Figure 17.1 Observed globally averaged land and ocean surface temperature anomaly, 1850
to 2012 relative to the period 1961 to 1990, for two data sets: (a) annual averages and (b)
decadal averages with estimated uncertainty for black line data set. (From IPCC, 2013)

Intergovernmental Panel on Climate Change (IPCC), some major findings should
concern anyone interested in our planet and its future. The IPCC was established
by the World Meteorological Organization and the United Nations Environmental
Programme to assess scientific, technical, and socioeconomic information relevant for
understanding climate change, its potential impacts, and options for adaptation and
mitigation. Some of the dominant conclusions of the panel, drafted in its most recent
multivolume reports and summaries (IPCC, 2013, 20I4a,b; The Royal Society and
The National Academy of Sciences, 2014) are summarized here:

■ The^ilobal average surface temperature has increased over the period 1880 to
2012 by about 0.83“ C. This trend is illustrated in Figure 17. 1. The temperate
increase was about 0.25°C more than that estimated by the IPCC (2001) for
the twentieth century (0.6°C). This temperature increase in the twentieth

Climate Change 565

century had also been determined to be the largest increase in the last 1,000
years.

■ Each of the last three decades has been successively warmer at Earth’s surface than
any preceding decade since 1850. In the Northern Hemisphere, 1983 to 2012
was likely the warmest 30-year period of the last 1,400 years.

■ The surface 75 m of oceans warmed by 0.11° C per decade over the period 1 992 to
2005. It is also likely that regions of high salinity, where evapotranspiration
(ET)»precipitation (P), have become more saline while humid regions, where
P>ET, have become fresher since the 1950s.

■ The average rate of ice loss from glaciers around the world, excluding glaciers on
the periphery of the ice sheets, was very likely 226 X 10'^ t ( Gt) yr^^over the period
1971 to 2009 and very likely 275 Gt yr~^ over the period 1993 to 2009. Over the
last two decades, the Greenland and Antarctic ice sheets have been losing
mass, glaciers continue to shrink worldwide, and Arctic sea ice and Northern
Hemisphere spring snow cover have continued to decrease.

■ Sea level has risen globally about 1.7 mm yr~^ (19 cm total) between 1901 and
2010 and at a much greater rate of 3.2 mm yr^^ from 1993 to 2010 (Fig. 17.2).
Glacial mass loss and ocean thermal expansion explains 75 percent of this
observed global sea level rise.

Causes of Climate Change

The cause of climate change is the increasing concentration of the so-called green-
house gases in the atmosphere, mostly caused by anthropocentric emissions. These
gases adsorb several wavelengths of long-wave radiation, causing Earth to be a little
warmer if the gas concentrations increase. The primary greenhouse gas is carbon diox-
ide (GO2), which is released through the burning of fossil fuels and also by cement
production. Atmospheric GO2 is estimated to have increased by over 30 percent since

satellite

Figure 17.2 Relative glebal mean sea level for 1955 to 2014 (Grey line reconstructed from
several data sets aligned to have the same value in 1993, the first year of satellite altimetry
data). (Ftom the Royal Society and the National Academy of Sciences, 2014)

566 Chapter 17 Wetlands and Climate Change

2010 2011 2012 2013 2014 2015

Year

Figure 17.3 (a) Concentration of CO2 in atmosphere at Mauna Loa Observatory in Hawaii
for 1958 through mid-2014. (b) Details of iast four years of seasonal CO2 fluctuations. Peaks
are at the beginning of the Northern Hemisphere growing season, after which photosynthe-
sis reduces the CO2 until the end of the growing season. (Data from Scripps Institution of
Oceanography and NOAA Earth System Research Laboratory)

the mid-eighteenth century. The longest record of continuous monitoring of CO2 in
the atmosphere is at Mauna Loa, Hawaii, started by C. David Keeling of the Scripps
Institution of Oceanography in March 1958 at a National Oceanic and Atmospheric
Administration (NOAA) facility (Fig. 17.3a). CO2 increased at a rate of 2.2 ppmyr“^
over 2009 to 2013, more than double the rate at which it was increasing in the
1960s. Monthly average concentrations reached 400 ppm during the spring of 2014
(Fig. 17.3b).

There has been much discussion about sources of CO2 besides fossil fuel burning,
such as tropical forest deforestation and burning. IPCC (2013) estimates that fossil
fuel combustion and cement production combined have released 375 Pg (= petagram
= Gt = 10^^ g) of CO2 as carbon to the atmosphere while deforestation and other
land-use changes are estimated to have released 180 Pg to the atmosphere. Fossil fuel
consumption continues to rise, from 6.7 Pg/yr in the mid-2000s to 10 Pg/yr in 2013.
The second most important greenhouse gas is actually water vapor, but it is not known
to have any trend or change. It is one of the most abundant gases in the troposphere.
When water vapor and other aerosols condense, they have a net negative radiative
forcing on the atmosphere, offsetting a major portion of the global mean radiative
forcing from other greenhouse gases (IPCC, 2013).

The third most important greenhouse gas is methane (CH4), which has been
estimated to have more than doubled in concentration, from about 720 parts per
billion (ppb) in preindustrial times to about 1,803 ppb in 2011. Before about 1980,
CH4 was assumed to be a stable concentration in the atmosphere, but it increased by
13 percent between 1978 and 1999 alone (Whalen, 2005). Wetlands were described
in Chapter 6: “Wetland Biogeochemistry” as being sources of CH4 gas, and that will
be put in context with other sources and sinks later in this chapter. What should be
clear is that if one argues that Earth has lost half of the world’s wetlands as a result of
human activity over the last 100 years when CH4Concentrations are increasing, there

Wetlands in the Global Carbon Cycle 567

is a disconnect. If wetlands were the major source of CH4, we would have seen a
decrease in CH4in the atmosphere over the last 100 years.

A fourth important greenhouse gas, nitrous oxide (N2O), also comes from wet-
lands as a result of nitrification and especially denitrification (see Chapter 6). While
N2O is a normal product of denitrification, it is usually a small percentage of denitri-
fication products, with most of nitrates converted to dinitrogen (N2) gas. N2O has
increased by about 20 percent in the atmosphere since preindustrial times.

Wetlands in the Global Carbon Cycle

Although soil carbon in wetland soils is recognized as an important component of
global carbon budgets and future climate change scenarios, very little work has been
done to consider the role of wetlands, particularly those in temperate and tropical
regions of the world, in the global carbon cycle. A carbon budget for the world, with
wetlands included to show their relative contributions, is shown in Figure 17.4. Fol-
lowing, we discuss the role of wetlands in this carbon budget in terms of carbon storage
in peat, carbon sequestration through peat and organic soil development, and CH4
emissions. This budget is a major revision from the budget published in the fourth edi-
tion of Wetlands (Mitsch and Gosselink, 2007). The major changes are a significantly
higher carbon sequestration of 1 Pg yr~^ (=1000 Tg (teragram) yr“^ = 10^® g yr“^)
estimated for the world’s wetlands, based on new data from several wetlands around

Figure 17.4 Global carbon budget with estimated role of wetlands in the carbon cycle.
Fluxes are in Pg/yr; storages are in Pg. Pg = 10^^ g. (CH4 emissions from wetlands and
rice paddies from Bloom et al., 2010; terrestrial ecosystem and fossil fuel inputs to CO2
from IPCC, 2013; carbon sequestration by wetlands from Mitsch et al., 2013)

568 Chapter 17 Wetlands and Climate Change

the world (Mitsch et al., 2013), and a continual increase in carbon emissions from
fossil fuel combustion from a mid-2000s estimate of 6.3 Pg/yr to the current rate of
10 Pg/yr, a 60 percent increase in emissions between two editions of this textbook.

Peat Storage and a Global Carbon Budget

Peat deposits in the world’s wetlands, particularly in boreal and tropical regions, are
substantial storages of carbon (C) in the lithosphere. Of the total storage of C in
Earth soils of 1,400 to 2,500 Pg-C (Pg= 10^^ g), anywhere from 20 to 30 percent
is stored in wetlands (Mitsch and Wu, 1995; Roulet, 2000; Hadi et ah, 2005; Lai,
2008). These peat deposits, if disturbed, however, could contribute significantly to
worldwide atmospheric CO2 levels, depending on the balance between draining and
oxidation of the peat deposits and their formation in active wetlands.

Carbon Sequestration

Many studies have now estimated carbon sequestration in wetlands in a variety of
temperate and tropical settings to augment the frequent estimates that already existed
for northern peatlands (Table 17.1).

For northern peatlands, the vertical accumulation rate of between 20 and
200 cm/1,000 yr (see Chapter 12: “Peatlands”) usually results in carbon accu-
mulation in the range of 10 to 50g-C m~^ yr“h This is a typical range of carbon
accumulation in peatiands (Table 17.1). A reasonable average of 29g-C m”^ yr“^
was found in a review of the literature for eight recent peatlands around the world
where carbon sequestration was measured (Table 17.1).

Most of the rates for carbon sequestration in tropical/subtropical wetlands and
for coastal mangroves and salt marshes are in the range of 150 to 250 g-C m"^ yr“^
(Table 17.1). Carbon sequestration by coastal wetlands (salt marshes, mangroves,
sea grasses) now has enormous international support and recognition, partially
because it is referred to in the literature and popular press as blue carbon (Mcleod
et al., 2011; Vaidyanathan, 2011; World Wildlife Fund, 2012; also see http://
thebluecarboninitiative.org/). The tropical wetlands included some high rates of
carbon sequestration such as seen for Cyprrwr wetlands in Uganda (Saunders et ah,
2007) but also relatively low rates of carbon sequestration in seasonally flooded
wetlands in Costa Rica and Botswana (Bernal and Mitsch, 2013b). In a study of
long-term accumulation in the tropics. Page et al. (2004) investigated a 9.5-m core of
peat from a tropical peatland in Kalimantan, Indonesia, and found an average carbon
sequestration of the core of 56 g-C m~^ yr”'^ over a 24,000-year period and a carbon
sequestration rate of 94 g-C m”^ yr“^ for the past 500 years in the upper meter of the
core (Table 17.1). The accumulation of peat in tropical wetlands may be due more
to the slow decomposition of recalcitrant lignin in roots and woody material under
constant high water rather than to high productivity of these systems (Chimner and
Ewel, 2005). The lower rates of carbon sequestration in seasonally flooded tropical
wetlands are probably due to the high temperatures year-round, especially in the dry
season, when some of the carbon is oxidized, or in some cases the presence of fire.

Table 17.1 Carbon Sequestration in wetiands (updated from Mitsch et ai., 2013) m~^

Wetland Type

Carbon Sequestration,
g-C year^

Reference

Northern Peatlands

Boreal peatlands

29 ± 13 (n = 8)

Mitsch et al. (2013)

Boreal peatlands

15-26

Turunen et al. (2002)

Temperate peatlands

10-46

Turunen et al. (2002)

Russian tundra peatlands

-8-38

Fleikkinen et al. (2002)

Coastai Wetiands

Mangroves, North America

180

Chmura et al. (2003)

Salt marshes, North America

220

Chmura et al. (2003)

Tidal freshwater wetlands, North America

140 ± 20

Craft (2007); Craft et al. (2009)

Brackish marshes. North America

240 ± 30

Craft (2007); Craft et al. (2009)

Salt marshes. North America

190 ± 40

Craft (2007); Craft et al. (2009)

Mangrove swamps, S.E. Asia

90 - 230

Suratman (2008)

Coastal wetlands, S.E. Australia,

105 - 137

Flowe et al. (2009)

Undisturbed sites

Coastal wetlands, S.E. Australia,

64-89

Flowe et al. (2009)

Disturbed sites

Mangroves (global)

160 ± 40

Breithaupt et al. (2012)

Mangroves (global)

226 ± 39

Mcleod etal. (2011)

Ttopical/Subtropicai Freshwater
Wetiands

Tropical/subtropical wetlands

194 ± 56 (n = 6)

Mitsch et al. (2013)

Florida Everglades, general

86-387

Reddy et al. (1993)

Tropical freshwater wetland, Indonesia

56 (for 24,000-year core)

Page et al. (2004)

Tropical freshwater wetland, Indonesia

94 (for last 500-year core)

Page et al. (2004)

Cyperus wetland In Uganda

480

Saunders et al. (2007)

Cypress (Taxodium) swamp, Florida

122

Craft et al. (2008)

Cypress {Taxodium) swamp, Georgia

Craft etal. (2008)

Everglades (Cladium) marsh, Florida

19-46

Craft et al. (2008)

Tropical flovrthrough swamp, Costa Rica

222-465 (ave = 306 for 3 sites)

Bernal and Mitsch (2013b)

Tropical forest basin wetland, Costa Rica

61-131 (ave = 84 for 3 sites)

Bernal and Mitsch (2013b)

Seasonally dry tropical floodplain

80-89 (ave = 84 for 3 sites)

Bernal and Mitsch (2013b)

wetland. Cost Rica

Seasonally flooded tropical floodplain

33-53 (ave = 42 for 3 sites)

Bernal and Mitsch (2013b)

wetland, Botswana

Florida Everglades — cypress

Villa and Mitsch (2015)

strand/swamp

Florida Everglades — pond cypress

Villa and Mitsch (2015)

Florida Everglades — wet prairie

Villa and Mitsch (2015)

Florida Everglades — upland pine flatwood

Villa and Mitsch (2015)

Temperate Freshwater Wetiands

Temperate wetlands

278 ± 42 (n = 7)

Mitsch et al. (2013)

Temperate flowthrough wetlands,

140 ± 16 (n = 3)

Bernal and Mitsch (2012)

northern Ohio

Depressional wetlands, Ohio

317 ± 93 (n = 3)

Bernal and Mitsch (2012)

Reed {Phragmites) marsh, Denmark

504

Brix et al. (2001)

{continued)

569

570 Chapter 17 Wetlands and Climate Change

Table 17.1 (Continued)

Carbon Sequestration,

Wetland Type

g-C m~2 year~^

Reference

Created and Restored Wetlands

Prairie pothole wetlands, North America
Restored (semipermanently flooded)

305

Euliss et al. (2006)

Reference wetland

Abandoned peat meadow, Netherlands
Created temperate riverine flowthrough

280

Hendriks et al. (2007)

marshes, Ohio
10 years old

181-193

Anderson and Mitsch (2006)

15 years old

219-267

Bernal and Mitsch (2013a)

Reference wetland

140

Bernal and Mitsch (2013a)

Temperate freshwater wetlands showed some of the highest rates of carbon
sequestration of any of the three climates investigated by Mitsch et al. (2013). Car-
bon sequestration in temperate-zone wetlands range from 230 to 320 to g-C m”^ yr“^
(Table 17.1). Brix et al. (2001) estimated a high rate of more than 500 g-C m~^ yr“^
in a productive Phra^mites marsh in Denmark.

Created and restored wetlands might be the best opportunity tor carbon seques-
tration. A carbon sequestration rate of 180 to 190 g-C m~^ yr”'^ for two created
wetland basins in Ohio (Anderson and Mitsch, 2006) 10 years after the wetlands
were created increased to 220 to 270 g-C yr“^ by the time the wetlands were
15 years old (Bernal and Mitsch, 2013a; Figure 17.5). About one -fourth of that car-
bon sequestration was as inorganic carbon, precipitated as calcite/calcium carbonate
(CaC03) due to high productivities in the water column. Euliss et al. (2006) com-
pared the carbon sequestration in several wedands that had been restored for more
than a decade in the prairie pothole wetlands of North America and found 305 g-C
m~^ yr“\ one of the highest numbers in Table 17.1. This is not surprising, because
restoration in these cases meant reflooding agricultural land, allowing organic carbon
to once again build up in the soil. For comparison, Euliss et al. (2006) estimated an
accumulation rate in reference (natural) marshes in the region of 83 g-C m”^ yr“^
based on average sedimentation rates of 2 mm/yr.

Methane Emissions

Wetlands are estimated to emit about 20 to 25 percent of current global CH4 emis-
sions or about 115 to 170 Tg-CH4 yr”^ (Tg= 10^^ g; Table 17.2). Thus, in climate
change discussions concerning wetlands, these “natural emissions” often receive the
most attention. Rice paddies, which are essentially domestic wetlands, account for
another 60 to 80 Tg-CH4 yr“h Other anthropogenic sources account for most of
the rest. CH4 emissions are a concern because CH4 is estimated to be 25 times more
effective as a greenhouse gas on a molecular basis than is CO2 after 100 years.

Year

Figure 17.5 Total soil carbon accumulation in two primary-succession, flowthrough, 1-ha
created wetlands in central Ohio over 15 years (1994 to 2009). 1995 carbon data are from
Nairn (1996); 2004 data are from Anderson et al. (2005) and Anderson and Mitsch (2006);
and 2009 data are from Bernai and Mitsch, 2013a). “Pianted wetiand” was pianted with 2,500
individuai plants representing 13 native plant species in May 1994; the “Naturally colonizing
wetland” remained as an unplanted centrel. Each wetland had identical hydrelogy for the
15 years (Mitsch et ai., 2012). (From Bernal and Mitsch, 2013a)

Table 17.2 Estimates of annuai fluxes of methane from wetlands and ether seurces,
Tg-CH,j/yr^

Sources

Megonigal et al. (2004)

Whalen (2005)

Bloom et al. (2010)

Natural wetlands

115

145

170

Tropics

Northern latitude

Others

Other Natural Sources'’

Anthropogenic

Rice Paddies

Other"

315

330

TOTAL SOURCES

535

600

=Tg = 10i2g

'’Other natural sources include termites, ocean, freshwater, and geological sources.

"Other anthropogenic sources include fossil fuels, landfills, domestic wastewater treatment, animal waste,
enteric fermentation (ruminants), and biomass burning.

571

572 Chapter 17 Wetlands and Climate Change

Tropical wetlands have been described recendy as more important than originally
thought for methane emissions (IPCC, 2013). Bloom (2010) suggests that 58 percent
(132 Tg-CH4 yr“^) of the total methane emissions from wedands and rice paddies
(227 Tg-CH4 yr”'^; see Table 17.2) comes from the tropics. Sjdgersten et al. (2014)
used a web analysis of current literature to esdmate 90 ± 77 Tg-CH4 yr”^ of methane
emissions from tropical wedands. They suggest that the methane emissions in the
tropics are greater from mineral soil wetlands than organic soil wedands.

CH4 emissions are actually the result of two compedng processes going on at the
same dme by microbial communities (see p. 197, “Methanogenesis,” and “Methane
Oxidation,” in Chapter 6: “Wetland Biogeochemistry”) (Fig. 17.6). The degradation
of organic matter by aerobic respiration is fairly efficient in terms of energy trans-
fer. Because of the anoxic nature of wedand soils, anaerobic processes, which are less

3 r Methane

Oxidized

V oxidation

soil or water

^ CH,

AnoxK

SCMl

Methanogenesis

Methane

Methano-

genesis

Oxidized
soil or
water

Anoxic

soil

Figure 17.6 Conceptual model ef CH^ emissions, ebullition, and CH^ oxidation in wetland
soils and plant vascular system. (Frem Conrad, 1993 and Whalen, 2005)

Wetlands in the Global Carbon Cycle 573

efficient in terms of energy transfer, occur in close proximity to aerobic processes.
Methcmo^enesis occurs when microbes called methanogens use CO2 as an electron
acceptor for the production of gaseous CH4 or, alternatively, use a low-weight organic
compound, such as one from a methyl group. CH4 production requires extremely
reduced conditions, with a redox potential of less than -200 mv, after other termi-
nal electron acceptors oxygen (O2), nitrates (NOg"), and sulfates (S04~) have been
reduced.

Conversely, nonflooded upland soils (e.g., forests, grasslands, arable land) are
regarded as the major biological sink of atmospheric CH4 (the major sink overall is tro-
pospheric photochemistry). Obligate aerobic methanotrophic bacteria use molecular
oxygen to oxidize CH4 to CO2 and cellular carbon. The consumption of atmospheric
CH4 is the result of two physiologically distinct microbial groups: (1) the methan-
otrophs, which have a membrane -bound enzyme system, and (2) an autotrophic nitri-
fier community. Methanotrophs are estimated to consume about 30 Tg CH4 yr“^
(Whalen, 2005).

CH4 production is much higher in the freshwater wetlands than from saltwater
wetlands. A major reason for low CH4 emissions from saltwater wetlands is the high
concentration of sulfates in seawater relative to freshwater that competes with car-
bon for oxidizable substrate (see “Carbon-Sulfur Interactions” in Chapter 6). CH4
emissions from studies of various freshwater wetlands around the world show have
a considerable range (Table 17.3) and measurements at a given wetland are rarely
normally distributed. Ebullition (see Fig. 17.6) is frequent yet hard to measure with
enough frequency. In a word, it is extraordinarily difficult to obtain accurate and
repeatable CH4 emission measurements from wetlands.

Most early CH4 emission studies were done in northern peatlands (bogs and fens)
in cold climates. Moore and Roulet (1995) suggested that most annual CH4 emission
flux measurements in Canada are less than 10 g CH4 m~^ yr”^ with the primary con-
trolling mechanisms being soil temperature, water table position, or a combination of
both. We estimate from recent studies using modern field and laboratory methods that
the general range of CH4 emissions from boreal wedands is from 15 to 25 g-C m~^
yr“^ (Table 17.3). An early estimate of CH4 emissions by Gorham (1991) that has
been used for determining the global contributions of northern peatlands is 28 g-C
m~^ yr”h In general, CH4 emissions from bogs are much lower than those from
the more mineral-rich fens. Aselmann and Crutzen (1989) assumed rates of CH4
emissions in the order of increasing emissions is bogs<fens<swamps<marshes<rice
paddies. Temperate wetlands emit CH4 generally in the range of 40 to 75 g-C
yr“^ (Table 17.3) although numbers are often quite variable.

In an interesting comparison of created versus natural wetlands in temperate
climates, Nahlik and Mitsch (2010) found that CH4 emissions in a reference
natural flowthrough wetland in Ohio were almost twice the emission rates found
in 15-year-old created flowthough marshes in Ohio (57 vs. 30 g-C m”^ year~^;
Table 17.3). This suggests that created and restored wetland CH4 emissions, even 15
years after the wetlands are created, may not be nearly at rates comparable to natural
wetlands yet.

574 Chapter 17 Wetlands and Climate Change

In those same created wetlands in Ohio a few years before, CH4 emissions were
compared during a year when both wetlands were pulsed with six hydrologic pulses
with pumped water floods, one each in months January through June 2004, the nor-
mal wet season with the following year (2005) when pumped water was steady flow all
year long. CH4 emissions during the seasonally pulsed year were considerably lower
in continuously flooded zones in the flood pulsing year than in the steady flowing year
(Altor and Mitsch, 2008) (Fig. 17.3). There was also a considerable difference in CH4
emissions in both years between the continuously flooded zones and the intermit-
tently flooded edge zones of the wedands (Table 17.3). These results have significant

Table 17.3 Methane emissions from freshwater wetlands (Updated from Mitsch et al., 2013)

Methane Emissions,

Climate and Wetland Type g-C Reference

Boreal Wetlands

Peatlands, general
Canadian peatlands
Russian peatlands

Ttopical/Subtropical Freshwater Wetlands

Tropical/subtropical wetlands
Amazon basin, Brazil
Amazon basin, Brazil
Orinoco floodplain, Veneuzela
Tropical flowthrough wetland, Costa Rica
Disturbed tropical floodplain wetland, Costa Rica
Tropical rain forest basin wetland, Costa Rica
Tropical seasonally flooded marsh, Botswana
Subtropical cypress strand, southwest Florida (4 wetland
communities)

Temperate Freshwater Wetlands

Temperate wetlands
Australian billabong
Temperate forested wetlands
Freshwater marsh, Virginia
Louisiana freshwater marshes
Louisiana bottomland hardwood forest
Spring-fed wetlands, Mississippi
Flowthrough wetlands, Ohio (experimental pulsing and
steady flow years in edge zones)

Flowthrough wetlands, Ohio (experimental pulsing year in
continuously flooded zones)

Flowthrough wetlands, Ohio (experimental steady flow year
in continuously flooded zones)

Created temperate marshes, Ohio
Reference flowthrough wetlands, Ohio

19 ± 7 (n =8)

Mitsch et al. (2013)

<7.5

Moore and Roulet (1995)

-1.2 - 12

Heikkinen et al. (2002)

119 ± 40 (n=6)

Mitsch et al. (2013)

40 - 215

Devol et al. (1988)

Melack et al. (2004)

Smith et al. (2000)

33 ± 5

Nahlik and Mitsch (2011)

263 ± 64

Nahlik and Mitsch (2011)

220 ± 64

Nahlik and Mitsch (2011)

72 ± 8

Mitsch et al. (2013)

1-49

Villa and Mitsch (2014)

58 ± 15 (n=7)

Mitsch et al. (2013)

12 - 22

Sorrell and Boon (1992)

Bartlett and Harriss (1993)

Whiting and Chanton (2001)

3 - 225

Delaune and Pezeshki (2003)

Yu et al. (2008)

Koh et al. (2009)

19 ± 6

Altor and Mitsch (2008)

49 ± 9

Altor and Mitsch (2008)

97 ± 19

Altor and Mitsch (2008)

30 ± 14

Nahlik and Mitsch (2010)

57 ± 18

Nahlik and Mitsch (2010)

Wetlands in the Global Carbon Cycle 575

Figure 17.7 Mean methane flux rates from experimental wetlands in created riverine wet-
iands in central Ohio during a flood pulsed year (2004) and a steady flow year (2005). Diffei^
ent letters represent a significant difference (p <0.05) between wetland zones, seasons, or
years. Bars represent standard error. (From Altor and Mitsch, 2008)

implications for keeping rivers and riverine wetlands free flowing with periodic floods,
as opposed to flow regulation where water levels and flow rates are maintained at
constant rates. Floods are good for minimizing CH4 emissions.

Comparing Apples and Oranges: The Net Balance of Methane
Production and Carbon Sequestration of Wetlands

There is a lot of confusion on the part of wetland conservationists, ecological
engineers who are creating and restoring wetlands, and climatologists as to
where wetlands fit into climate change. On one hand, wetlands are creating
a greenhouse gas, CH4 (and have been doing so for the ages), but on the
other hand, wetlands of the world are sequestering carbon, some at significant
rates. In fact, some of the fossil fuels that are now running our economy come
from the organic carbon sequestered by swamps. So are wetlands good or bad
for climate change?

Mitsch et al. (2013) developed a dynamic carbon model (Fig. 17.8)
that included both soil carbon sequestration and CH4 emissions to
investigate this question. The model featured two carbon exchanges with the

Figure 17.8 Wetland carbon simulation model designed to estimate the net effects
of carbon sequestration and CH^ emissions over time, assuming a global warming
potential (GWP) of methane relative to CO2 and linear atmospheric decay of methane.
(GPP = gross primary productivity; Rp = plant respiration; R3 = soil respiration;
net carbon sequestration; F„,g = methane emissions). (From Mitsch et al., 2013)

atmosphere — CH4 emissions from the wetland to the atmosphere and CO2
exchange to the wetland from the atmosphere. Model parameters include a
half-life of seven years for CH4 and a global warming potential (GWP) for CH4.
CH4 emission and carbon sequestration data from 16 natural wetlands from
around the world were used as inputs for the model simulations. The CO2
equivalent is determined as:

C02eq = CO2 -t (GWPm x Mch4) (17.1)

where CO2 = atmospheric carbon dioxide, g-C02 m~^

Mch4 = atmospheric methane, g-CH4

GWP|y| = global warming potential for methane = 25 for 100 years

576

Wetlands in the Global Carbon Cycle 577

Model simulations showed that most of the 16 wetlands become net sinks
of radiative forcing well within the 100 years. This is because the impact of
CH4 emissions is temporary in the atmosphere; CH4 eventually decays to CO2
and is “trumped” by the permanent burial of carbon in the wetland soil. In this
set of simulations, only 2 of the 16 wetlands remain radiative sources; both
were Russian peatlands that were already CO2 sources because they had been
drained. If a wetland is a CO2 source, then it will always be a source of radiative
forcing. This model results suggests that if the natural hydrology of a wetland
is intact and the wetland is sequestering some C02from the atmosphere, it
will, with little question, be a net sink of radiative forcing and thus good for
the climate.

Climate Change Feedbacks

One of the interesting questions about the vast storages of peat in northern climes
related to the potential positive feedback to climate change that could occur. Because
there is significantly more carbon stored in the world’s soils than in the atmosphere
(see Fig. 17.4), there is the potential that if the climate were to warm and accelerate
decomposition of peatlands, then these peatlands would become an additional major
source of carbon, through aerobic respiration and possibly fires, to the atmosphere.
Davidson and Janssens (2006) summarize the comparison of uplands, which have
good drainage and aeration and are therefore less prone to having large releases of
C02in the event of warming, to peadands, where drainage is poor and soils are anaero-
bic. They describe peatland soils as enormously vulnerable to climate change compared
to upland soils (Table 17.4), even though peadand soils make up a reladvely small
percentage of Earth’s landscape. The release of 100 petagrams of carbon (Pg-C) from
peadands by the year 2100 would mean that for several years, carbon would be released
at rates comparable to those currently caused by fossil fuels. If peadand producdvity
were to increase with the increase in temperature, it could offset this positive feedback
and even lead to a negadve feedback, where more carbon is sequestered than released.

Table 17.4 Below-ground carbon stocks in the world and their vulnerabilities to loss by 2100
due to global warming

Carbon Pool

Carbon Size, Pg-C

Potential Loss by 2100 from Global Warming

Upland soil Inventory (3 m depth)

2,300

0-40

Peatlands (3 m depth)

450

100

Permafrost

400

100

Source: Davidson and Janssens (2006)

578 Chapter 17 Wetlands and Climate Change

Christensen (1991) predicted that, as a result of a 5 percent global warming, the
tundra would change from being a net sink of CO2 to a net source of up to 1 .25 Pg/yr
carbon because of a combination of thermokarst erosion, deepening of the active layer
in permafrost areas, lowering of the water table, and higher temperatures. Tarnocai
(2006) was more direct and predicted severe degradation of the frozen peadands in
the subarctic and northern boreal Canada and severe drying in the southern boreal
regions as well, but a scenario of 3° to 5°C increase in air temperature and 5° to
7°C increase over the oceans by the end of the twenty-first century. The affected area
represents about 50 percent of all the organic carbon mass occurring in all Canadian
wetlands.

In general, both the increase in temperature and the changes in water levels are
important variables in the production of CH4 and CO2 from wetlands, but their rela-
tive importance for CH4 generation is poorly understood. Using a model with inputs
of a 3°C rise in temperature and a decrease in the water table between 14 and 22 cm for
a subarctic fen, Roulet et al. ( 1992) estimated that the increased temperature raised the
CH4 flux between 5 and 40 percent, but the lowered water table decreased the CH4
flux by 74 to 81 percent. This decrease in CH4 flux in drier conditions was caused by a
decrease in the zone of active methanogenesis and by an increase in CH4 oxidation in
the aerobic layer. Thus, the influence of global temperature rise would depend locally
on the temperature increase relative to the induced change in the moisture regime.

Carbon Budgets

Carbon budgets for peatlands have drawn a great deal of interest, given the
importance of these ecosystems in global carbon dynamics. A carbon budget
for individual created wetland basins was already presented in Chapter 6.
It is accepted that boreal peatlands were once carbon sinks, but there is
little consensus that they are contemporary sinks. Carbon budgets have
been developed for small peatlands (Carroll and Crill, 1997; Waddington
and Roulet, 1997) and for substantial-size peatland-dominated watersheds
(Rivers et al., 1998). In the latter, a l,500-km2 watershed In the Lake Agassiz
peatlands In Minnesota illustrated that the peat watershed had a net carbon
storage of 12.7 g-C m“^ yr“^ but that there was a tenuous balance between
the watershed being a source and a sink of carbon (Fig. 17.9). Inflows
of carbon are groundwater, precipitation, and net community productivity,
while outflows are groundwater and surface flow and outgassing of CH4. It
was estimated from a companion study (Glaser et al., 1997) that peat is
accumulating at a rate of Imm/yr (100 cm/1000 yr). This budget illustrates
the importance of accurate hydrologic measurements as well as biological
productivity measurements in determining accurate carbon budgets for
wetlands and wetland landscapes.

Effects of Climate Change on Wetlands 579

Diffusive and advective

Dissolved inorganic carbon CH^ loss and gain

Figure 17.9 Carbon budget of the 1,500-km^ Rapid River watershed in the Lake Agas-
siz peatiand basin of northern Minnesota. Fluxes are in g-C m~^ yr~^. (After Rivers
et al., 1998)

Effects of Climate Change on Wetlands

Wetlands may be key ecosystems for mitigating the effects of fossil fuel emissions on
climate. Conversely, sea-level and temperature changes may have significant impacts
on coastal and inland wetlands.

Coastal Wetlands

One of the major impacts of possible climate changes on wetlands is the effect that
sea-level rise will have on coastal wedands. Estimates of sea-level rise over the next
century range from 50 to 200 cm. (Fig. 17.2 shows the current rate of sea level rise of
32 cm/century.) It has been estimated that if sea level were to rise by 100 cm, half of
the wetlands designated by the Ramsar Convention as wetlands of international impor-
tance would be threatened (Nicholls, 2004). The regions where wetlands are most at
risk, even for a 44-cm rise in sea level by 2080, are shown in Figure 17.10. If the rise
in sea level is not accompanied by equivalent vertical accretion of marsh sediments,
then coastal marshes will gradually disintegrate as a result of increased inundation,
erosion, and saltwater intrusion. Because much of the coastline of the world is devel-
oped, efforts to protect dry upland from inundation by the construction of bulkheads
or dikes will exacerbate the problem. In essence, the wetlands will be trapped between
the rising sea and the protected dry land, a situation that has already occurred over

580 Chapter 17 Wetlands and Climate Change

Figure 17.10 Coastal wetland areas most vulnerable to a sea-level rise of 44 cm by 2080.
(Ftom IPCC, 2001)

the centuries in the Netherlands and China. This effect has been termed the coastal
squeeze of sea-level rise (Nicholls, 2004). Even in the absence of bulkheads in most
of our regions where coastal wedands exist, “the slope above the wetland is steeper
than that of the wetlands; so a rise in sea level causes a net loss of wetland acreage”
(Titus, 1991).

Estimates of the loss of coastal wetlands in the United States vary, with much of
the variability dependent on the assumed sea-level rise and the degree to which dry
land is protected at all cost (Table 17.5). If there is no shoreline protection, a sea-level
rise of 1 m could reduce coastal wetlands by 26 to 66 percent. If the policy were to
protect all dry land, then the estimated loss of wetlands increases dramatically to 50
to 82 percent. How well these figures can be extrapolated to the rest of the world is
unclear. In long-developed coastlines, such as those of Europe and the Far East, the
losses would probably be less.

Table 17.5 Estimated percentage coastal wetland loss in the United States with
sea-level rise

Sea-Level Rise

0.5 m

If no shores are protected

If densely developed dry land is protected

If all dry land is protected

17-43%

20-45%

38-61%

26-66%

29-69%

50-82%

29-76%

33-80%

66-90%

Source: Titus (1991)

Effects of Climate Change on Wetlands 581

The Mississippi River Delta in Louisiana may be a model for seeing the effects of
global sea-level rise on coastal wetlands. Here, the “apparent” sea-level rise is already
1 m/100 yr (1 cm/yr), primarily because of sediment subsidence rather than actual
sea-level rise. In this delta marsh, vertical accretion is not keeping up with subsidence,
in part because the Mississippi River is carrying only about 20 percent of the sediment
load it did in 1850 (Kesel and Reed, 1995) and its flow is contained within levees, so
riverborne sediments no longer reach the wedands during spring floods. As a result,
this region has the highest rate of wedand loss in the United States. Day et al. (2005)
describe the ramificadons of global climate change on restoradon efforts now under
way in the delta. With a sea-level rise of 30 to 50 cm by 2100 possible, the reladve
sea-level rise will increase from 1 cm/yr (caused mosdy by land subsidence) to 1.3 to
1.7 cm/yr, exacerbadng an already difficult situadon of wedand loss in the Louisiana
Delta. In addidon. Day et al. (2005) note that as a result of milder temperatures
already, mangrove swamps were beginning to replace their temperate -zone analog,
the salt marsh, in several locadons in the delta. This mangrove expansion is another
effect that would be expected in subtropical regions that were previously dominated
by salt marshes. Mangroves are valuable coastal ecosystems, as are salt marshes, but
the overall effects of this subsdtudon of ecosystems is unclear.

Management of Coastal Wetlands

There are few management possibilities for managing coastal wetlands in the face of
sea-level rise. Figure 17.11 shows two future condidons. In Future 1, the house is
protected with a bulkhead in the face of rising sea level, and the salt marsh is lost

Today

V current sea level
past sea level

Future 1

vegetation is lost
because of deep wiater

future sea level
current sea level

Future 2

vegetation moves inland and
establishes in intertidal zone

move house and
bulkhead inland

future sea level
current sea level

Figure 17.11 Coastal wetland management scenarios in the face of sea-level rise. Future 1
is without moving human habitation inland. Future 2 involves moving human activity inland
to allow room for the wetland to move inland. (Frem Titus, 1991)

582 Chapter 17 Wetlands and Climate Change

or “squeezed out.” In Future 2, the house is moved upland to accommodate the
wetland, which would begin to form if a gentle slope and adequate sediment sources
were available.

Future 2 models the wetlands of the Laurentian Great Lakes, which, for centuries,
were “wetlands on skateboards,” moving inland and lakeward with frequent (over
periods of decades) water-level changes in the lakes (Mitsch, 1992). With stabilization
of the coastline in the past century, diking the remaining wetlands along the Great
Lakes became necessary for their survival.

Day and Templet (1989) and Day et al. (2005) concluded, after extensive
investigation of the apparent sea-level rise in coastal Louisiana, that we can manage
coastal wetlands in periods of rising sea level through comprehensive, long-range
planning and through the application of the principles of ecological engineering
by using nature’s energies, such as upstream riverine sediments and fresh water,
vegetation productivity, winds, currents, and tides, as much as possible.

Inland Wetlands

In addition to the effects of climate change on coastal wedands through sea-level rise,
the change in climate, particularly temperature (Fig. 17.1), will probably affect the
function and distribution of inland wetlands. In the tundra, any melting of the per-
mafrost would result in the loss of wetlands. In boreal and temperate areas, climate
change would result in changing rainfall patterns, thus affecting runoff and ground-
water inflows to wetlands. In general, a decrease in precipitation or an increase in
evapotranspiration will result in less-frequent flooding of existing wetlands, although
the types of wetlands may not change. Greater precipitation patterns would increase
the length and depth of flooding of inland wetlands. Most susceptible to these effects
are depressional wetlands that have small watersheds and that are in regions between
arid and mesic climates, such as the prairie potholes of North America.

The impact of climate change on the Prairie Pothole Region (PPR) of North
America was investigated by Johnson et al. (2005). These wedands provide 50 to 80
percent of the continent’s duck population and are exactly on the edge between areas
to the east with abundant precipitation and arid climates to the west. By using a wet-
land simulation model, Johnson et al. (2005) were able to predict areas in the pothole
region that would have highly favorable water conditions for three climate scenarios:
(I) a 3°C temperature increase with no change in precipitation; (2) a 3°C tempera-
ture increase with a 20 percent increase in precipitation; and (3) a 3°C temperature
increase with a 20 percent decrease in precipitation (Fig. 17.12). Basically any temper-
ature increase coupled with precipitation decrease shifted the area favorable for ducks
to the east. Overall, the climate change would “diminish the benefits of wetland con-
servation in the central and western PPR. Simulations further indicate that restoration
of wetlands along wetter fringes of the PPR may be necessary to ameliorate potential
impacts of climate change on waterfowl populations” (Johnson et al., 2005).

Effects of Climate Change on Wetlands 583

Figure 17.12 Simulation results for locations of highly favorable water and caver condi-
tions in the Prairie Pethole Regien of North America for waterfowl breeding under optiens
(a) histeric; (b) a 3°C temperature increase with no change in precipitation; (c) a 3°C tem-
perature increase with a 20 percent increase in precipitation; and (d) a 3°C temperature
increase with a 20 percent decrease in precipitation. (From Johnson et al., 2005, reprinted
with permission)

Management of Inland Wetlands

Limited experimentation, especially in rice paddies, suggests some management alter-
natives that might be appropriate for inland wetlands, especially to reduce methane
emissions. Sass et al. ( 1992) measured the effects on methane emissions of four differ-
ent water management methods in some rice fields in Texas and found that temporary
drainage (midseason drainage and multiple aeration) decreased CH4 emission caused
by both increased CH4 consumption in the aerobic layer and decreased CH4 produc-
tion. Such management may be practical only in flat systems with sufficient control of
water levels.

Nutrient and compost management may also offer opportunities for reducing
methane emissions. There may be a relationship between the carbomnitrogen (C:N)
ratio of the organic matter in wetiands and CH4 emissions, although the trends are
not clear. Yagi and Minami (1990) found, in rice paddies in Japan, that compost with a
low C:N ratio (enriched in nitrogen) causes lower emissions of methane than uncom-
posted rice straw with a high C:N ratio. Conversely, Schutz et al. (1989) found high

584 Chapter 17 Wetlands and Climate Change

emissions in fields applied with compost. Because of its competition with methano-
genesis, enhancing sulfate reduction is often suggested as a management alternative
to reduce CH4 emissions. This has long been known as one of the primary reasons
that methanogenesis is lower in saltwater wetlands than in freshwater wetlands.

One of the easiest management approaches for minimizing CH4 emissions from
freshwater wedands is to allow the wedands to have their natural fluctuating hydroperi-
ods and, in some cases, a pulsing hydrology. Studies by Altor and Mitsch (2006, 2008 )
described above showed that a pulsing hydrology had CH4 emissions that were much
lower than those from permanently flooded sites.

We cannot esdmate, at present, with much certainty whether wedands are sig-
nihcant global carbon sources or sinks. Nevertheless, the opportunides for managing
CO2 and CH4 emissions in wetlands are not generally on a scale large enough to make
much difference to the global carbon balance.

Recommended Readings

Mitsch, W. J., B. Bernal, A. M. Nahlik, U. Mander, L. Zhang, C. J. Anderson, S. E.
Jorgensen, and H. Brix. 2013. Wedands, carbon, and climate change. Landscape
Ecology 2^-. 583-597.

The Royal Society and the Nadonal Academy of Sciences. 2014. Climate Change:
Evidence and Causes. An overview from the Royal Society and the US Nadonal
Academy of Sciences. London and Washington DC: Royal Society and US NAS.
Whalen, S. C. 2005. Biogeochemistry of methane exchange between natural wedands
and the atmosphere. Environmental Engineering Science 22: 73-94.

References

Altor, A. E., and W. J. Mitsch. 2006. Methane flux from created riparian marshes:
Reladonship to intermittent versus condnuous inundation and emergent macro-
phytes. Ecological Engineering 28: 224-234.

Altor, A. E., and W. J. Mitsch. 2008. Pulsing hydrology, methane emissions, and car-
bon dioxide fluxes in created marshes: A 2-year ecosystem study. Wetlands 28:
423-438.

Anderson, C. J., W. J. Mitsch, and R. W. Nairn. 2005. Temporal and spadal devel-
opment of surface soil condidons in two created riverine marshes. Journal of
Environmental Quality 34: 2072-2081.

Anderson, C. J., and W. J. Mitsch. 2006. Sediment, carbon, and nutrient accumuladon
at two 10-year-old created riverine marshes. Wetlands 26: 779-792.

Aselmann, 1., and P. J. Crutzen. 1989. Global distribudon of natural freshwater wet-
lands and rice paddies, their net primary producdvity, seasonality and possible
methane emissions. Journal of Atmospheric Chemistry 8: 307-358.

Bardett, K. B., and R. C. Harriss. 1993. Review and assessment of methane emissions
from wedands. Chemosphere 26: 261-320.

References 585

Bernal, B., and W. J. Mitsch. 2012. Comparing carbon sequestration in temperate
freshwater wetland communities. Global Change Biology 18: 1636-1647.

Bernal, B., and W. J. Mitsch. 2013a. Carbon sequestration in two created riverine
wetlands in the Midwestern United States. Journal of Environmental Quality 42:
1236-1244.

Bernal. B., and W. J. Mitsch. 2013b. Carbon sequestration in freshwater wetlands of
Costa Rica and Botswana. Bio^eochemistry. 115: 77-93.

Bloom, A. A., R 1. Palmer, A. Fraser, D. S. Reay, and C. Frankenberg. 2010.
Large-scale controls of methanogenesis inferred from methane and gravity
spaceborne data. Science 327: 322-325.

Breithaupt, J. L., J. M. Smoak, T. J. Smith, C. J. Sanders, and A. Hoare 2012, Organic
carbon burial rates in mangrove sediments: Strengthening the global budget.
Global Biojjeochem. Cycles 26: GB3011.

Brix, H., B. K. Sorrell, and B. Lorenzen. 2001. Are Phrajjmites-dommiLtcd wetlands
a net source or net sink of greenhouse gases.^ Aquatic Botany 69: 313-324.

Carroll, R, and P. Crill. 1997. Carbon balance of a temperate poor fen. Global Bio^eo-
chemical Cycles 11: 349-356.

Chimner, R. A., and K. C. Ewel. 2005. A tropical freshwater wetland: 11. Produc-
tion, decomposition, and peat formation. Wetlands Ecology and Management 13:
671-684.

Chmura, G. L., S. G. Anisfeld, D. R. Gaboon, and J. G. Lynch. 2003. Global carbon
sequestration in tidal, saline wetland soils. Global Biogeo chemical Cycles 17:1111.

Ghristensen, T. 1991. Arctic and sub-Arctic soil emissions: Possible implications for
global climate change. Polar Record 27: 205-210.

Gonrad, R. 1993. Gontrols of methane production in terrestrial ecosystems. In M. O.
Andreae and D. S. Schimel, eds. Exchange of Trace Gases between Terrestrial Ecosys-
tems and the Atmosphere. John Wiley & Sons, New York, pp. 39-58.

Graft, G. 2007. Freshwater input structures soil properties, vertical accretion,
and nutrient accumulation of Georgia and U.S. tidal marshes. Limnol-
ogy and Oceanography 52:1220-1230.

Graft, G., G. Washburn, and A. Parker. 2008. Latitudinal trends in organic carbon
accumulation in temperate freshwater peatlands. In J. Vymazal, ed. Wastewa-
ter Treat Plant Dynamics and Management in Constructed Natural Wetlands.
Springer, Dordrecht, the Netherlands, p. 23-31.

Graft, G., J. Glough, J. Ehman, J. Samantha, R. Park, P. Pennings, H. Guo, and
M. Machmuller. 2009. Forecasting the effects of sea-level rise on tidal marsh
ecosystem services. Erontiers in Ecology and the Environment 7\7Z-T&.

Davidson, E. A., and 1. A. Janssens. 2006. Temperature sensitivity of soil carbon
decomposition and feedbacks to climate change. Nature 440: 165-173.

Day, J. W., Jr., and P. H. Templet. 1989. Gonsequences of sea level rise: Implications
from the Mississippi Delta. Coastal Management 17: 241-257.

Day, J. W., Jr., J. Barras, E. Glairain, J. Johnston, D. Justic, G. P. Kemp, J.-Y. Ko,
R. Lane, W. J. Mitsch, G. Steyer, P. Templet, and A. Yahez-Arancibia. 2005.

586 Chapter 17 Wetlands and Climate Change

Implications of global climatic change and energy cost and availability for the
restoration of the Mississippi Delta. Ecological Engineering 24: 253-265.

Delaune, R. D., and S. Pezeshki. 2003. The role of soil organic carbon in maintain-
ing surface elevation in rapidly subsiding U.S. Gulf of Mexico coastal marshes.
Water, Air and Soil Pollution 3: 167-179.

Devol, A. H., J. E. Richey, W. A. Clark, S. L. King. 1988. Methane emissions to
the troposphere from the Amazon floodplain. Journal of Geophysical Research 93 :
1492-1583.

Euliss, N. H., R. A. Gleason, A. Olness, R. L. McDougal, H. R. Murkin, R. D.
Robarts, R. A. Bourbonniere, and B. G. Warner. 2006. North American prairie
wetlands are important nonforested land-based carbon storage sites. Science of the
Total Environment 361: 179-188.

Glaser, P. H., P. C. Bennett, D. I. Siegel, and E. A. Romanowicz. 1997. Palaeo-
reversals in groundwater flow and peatland development at Lost River, Minnesota,
USA. Holocene 6: 413M:21.

Gorham, E. 1991. Northern peatlands: Role in the carbon cycle and probable
responses to climatic warming. Ecological Applications 1: 182-195.

Hadi, A., K. Inubushi, Y. Furukawa, E. Purnomo, M. Rasmadi, and H. Tsuruta.
2005. Greenhouse gas emissions from tropical peatlands of Kalimantan, Indone-
sia. Nutrient Cycling in A^roecosystems 7 1\ 73-80.

Heikkinen J. E. P, V. Elsakov, and P. J. Martikainen. 2002. Carbon dioxide and
methane dynamics and annual carbon budget in tundra wedand in NE Europe,
Russia. Global Bio^eochemical Cycles 16. doi:10.1029/2002GB0001930

Hendriks, D. M. D., J. van Huissteden, A. J. Dolman, and M. K. van der Molen. 2007.
The full greenhouse gas balance of an abandoned peat meadow. Bio^eosciences 4:
411-424.

Howe, A. J., J. F. Rodriguez, and P. M. Saco. 2009. Surface evolution and carbon
sequestration in disturbed and undisturbed wedand soils of the Hunter Estuary,
southeast Australia. Estuarine, Coastal and Shelf Science 84: 75-83.

Intergovernmental Panel on Climate Change (IPPC). 2001. Climate Change 2001:
The Scientific Basis. Published for the Intergovernmental Panel on Climate
Change. Cambridge University Press, UK.

Intergovernmental Panel on Climate Change (IPCC). 2013. Climate Change 2013:
The Physical Science Basis. Contribution of Working Group I to the Fifth Assess-
ment Report of the Intergovernmental Panel on Climate Change, ed. T. F. Stocker,
D. Qin, G.-K. Plattner, M. Tignor, S. K. Allen, J. Boschung, A. Nauels, Y. Xia, V.
Bex, and P. M. Midgley. Cambridge University Press, Cambridge, UK. 1535 pp.

Intergovernmental Panel on Climate Change. 2014a. Climate Change 2014: Impacts,
Adaptation, and Vulnerability. Parts A and B. Cambridge University Press, Cam-
bridge, UK. 1132 pp and 688 pp.

Intergovernmental Panel on Climate Change, 2014b. Climate Change 2014: Mitiga-
tion of Climate Chanpie, Workinjj Group III Contribution to the IPCC 5th Assess-
ment Report. IPCC, http://www.ipcc.ch/report/ar5/wg3/. 99 pp.

References 587

Johnson, W. C., B. V. Millett, T. Gilmanov, R. A. Voldseth, G. R. Guntenspergen, and

D. E. Naugle. 2005. Vulnerability of northern prairie wetlands to climate change.
BioScience 55: 863-872.

Kesel, R., and D. J. Reed. 1995. Status and trends in Mississippi River sediment regime
and its role in Louisiana wetland development. In D. J. Reed, ed. Status and
Historical Trends of Hydrologic Modification, Reduction in Sediment Availabil-
ity, and Habitat Loss/Modification in the Barataria-Terrebonne Estuarine System.
BT/NEP Publication 20, Barataria-Terrebonne National Estuary Program, Thi-
bodaux, LA, pp. 80-98.

Koh, H.-S., G. A. Ochs, and K. Yu. 2009. Hydrologic gradient and vegetation controls
on GH4 and GO2 fluxes in a spring-fed forested wetland. Hydrobiologia 630:
271-286.

Lai, R. 2008. Garbon sequestration. Philosophical Transactions of the Royal Society B
363: 815-830.

Mcleod, E., G. L. Ghmura, S. Bouillon, R. Salm, M. Bjork, G. M. Duarte, G. E.
Lovelock, W. H. Schlesinger, and B. RSilliman. 201 1 . A blueprint for blue carbon:
toward an improved understanding of the role of vegetated coastal habitats in
sequestering GO2. Frontiers ofEcolojjy and the Environment 9: 552-560.
Megonigal, J. P, M. E. Hines, and P. T. Visscher. 2004. Anaerobic metabolism:
Linkages to trace gases and aerobic processes. In W. H. Schlesinger, ed. Bio^eo-
chemistry. Elsevier-Pergamon, Oxford, UK, pp. 317-424.

Melack, J. M., L. L. Hess, M. Gastil, B. R. Forsberg, S. K. Hamilton, I. B. T. Lima, and

E. M. L. M. Novo. 2004. Regionalization of methane emissions in the Amazon
Basin with microwave remote sensing. Global Change Biology 10:530-544.

Mitsch, W. J. 1992. Gombining ecosystem and landscape approaches to Great Lakes
wetlands. Journal of Great Lakes Research 18:552-570.

Mitsch, W. J., and X. Wu. 1995. Wetlands and global change. In R. Lai, J. Kimble,
E. Levine, and B. A. Stewart, eds. Advances in Soil Science, Soil Management, and
Greenhouse Effect. GRG Press/Lewis Publishers, Boca Raton, FL, pp. 205-230.
Mitsch, W. J., and J. G. Gosselink. 2007. Wetlands, 4th ed. John Wiley & Sons, Hobo-
ken, New Jersey, 582 pp.

Mitsch, W. J., L. Zhang, K. G. Stefanik, A. M. Nahlik, G. J. Anderson, B. Bernal,
M. Hernandez, and K. Song. 2012. Greating wedands: Primary succession, water
quality changes, and self-design over 15 years. BioScience 62: 237-250.

Mitsch, W. J., B. Bernal, A. M. Nahlik, U. Mander, L. Zhang, G. J. Anderson, S. E.
Jorgensen, and H. Brix. 2013. Wetlands, carbon, and climate change. Landscape
Ecology 28: 583-597.

Moore, T. R., and N. T. Roulet. 1995. Methane emissions from Ganadian peadands.
In R. Lai, J. Kimble, E. Levine, and B. A. Stewart, eds. Advances in Soil Science:
Soils and Global Change. GRG Press, Boca Raton, FL, pp. 153-164.

Nahlik, A. M., and W. J. Mitsch. 2010. Methane emissions from created riverine wet-
lands. Wetlands 30: 783-793. Erratum in Wetlands {2011) 31: 449-450.

588 Chapter 17 Wetlands and Climate Change

Nahlik, A. M., and W. J. Mitsch. 2011. Methane emissions from tropical freshwater
wetlands located in different climatic zones of Costa Rica. Global Chan£ie Biology
17: 1321-1334.

Nairn, R. W. 1996. Biogeochemistry of newly created riparian wetlands: Evaluation
of water quality changes and soil development. Ph.D. dissertation, Graduate Pro-
gram in Environmental Science, Ohio State University, Columbus.

Nicholls, R. J. 2004. Coastal flooding and wedand loss in the 21st century: Changes
under the SUES climate and socio-economic scenarios. Global Environmental
Change 14: 69-86.

Page, S. E., R. A. J. Wust, D. Weiss, J. O. Rieley, W. Shotyk, and S. H. Limin.
2004. A record of late Pleistocene and Holocene carbon accumulation and cli-
mate change from an equatorial peat bog (Kalimantan, Indonesia): Implications
for past, present, and future carbon dynamics. Journal of Quaternary Science 19:
625-635.

Reddy, K. R., R. D. DeLaune, W. F. DeBusk, and M. S. Koch. 1993. Long-term nutri-
ent accumulation rates in the Everglades. Soil Science Society of America Journal
57: 1147-1155.

Rivers, J. S., D. 1. Siegel, L. S. Chasar, J. P. Chanton, P. H. Glaser, N. T. Roulet, and
J. M. McKenzie. 1998. A stochastic appraisal of the annual carbon budget of a
large circumboreal peatland. Rapid River Watershed, northern Minnesota. Global
Bio^eochemical Cycles 12: 715-727.

Roulet, N. T. 2000. Peatlands, carbon storage, greenhouse gases, and the Kyoto Pro-
tocol: Prospects and significance for Canada. Wetlands 20: 605-615.

Roulet, N. T., T. R. Moore, J. Bubier, and P. Lafleur. 1992. Northern fens: CH4 flux
and climate change. Tellus 4:AB\ 100-105.

Sass, R. L., F. M. Fisher, Y. B. Wang, F. T. Turner, and M. F. Jund. 1992. Methane
emission from rice fields: The effect of floodwater management. Global Bio^eo-
chemical Cycles 6: 249-262.

Saunders M. J., M. B. Jones, and F. Kansiime. 2007. Carbon and water cycles in
tropical papyrus wetlands. Wetlands Ecology and Management 15: 489-498.

Schutz, H., A. Holzapfel-Pschorn, R. Conrad, H. Rennenberg, and W. Seiler. 1989.
A three year continuous record on the influence of daytime, season and fertilizer
treatment on methane emission rates from an Italian rice paddy held. Journal of
Geophysical Research 94: 16405-16416.

Sjogersten, S., C. R. Black, S. Evers, J. Hoyos-Santillan, E. L. Wright, and B. L.
Turner. 2014. Tropical wedands: A missing link in the global carbon cycle.^ Global
Biogeochem. Cycles 28. doi:10.1002/2014GB004844.

Smith, K.A.,K.E. Dobbie, B.C. Ball, et al. 2000. Oxidation of atmospheric methane
in Northern European soils, comparison with other ecosystems, and uncertainties
in the global terrestrial sink. Global Change Biology 6: 791-803.

Sorrell, B. K., and P. 1. Boon. 1992. Biogeochemistry of billabong sediments. 11.
Seasonal variations in methane production. Ereshwater Biology 27: 435-445.

References 589

Suratman, M. H. 2008. Carbon sequestration potential of mangroves in southeast
Asia. In F. Bravo, V. LeMay, R. Jandl, and K. von Gadow, eds., Mana^in^ For-
est Ecosystems. The Challen£;e of Climate Change. Springer Science, Dordrecht,
Netherlands, pp. 297-315.

Tarnocai, C. 2006. The effect of climate change on carbon in Canadian peatlands.
Global and Planetary Change 53: 222-232.

The Royal Society and the National Academy of Sciences. 2014. Climate Chan£ie:
Evidence and Causes. An overview from the Royal Society and the US National
Academy of Sciences, Royal Society and US NAS, London and Washington DC.

Titus, J. G. 199 1 . Greenhouse effect and coastal wetland policy: How Americans could
abandon an area the size of Massachusetts at minimum cost. Environmental Man-
agement 15: 39-58.

Turunen, J., E. Tomppo, K. Tolonen, and E. Reinkainen. 2002. Estimating carbon
accumulation rates of undrained mires in Finland: Application to boreal and sub-
arctic regions. The Holocene 12: 79-90.

Vaidyanathan, G. 2011. “Blue carbon” plan takes shape. Nature. doi:10. 1038/news
.2011.112

Villa, J. A. and W. J. Mitsch. 2014. Methane emissions from five wedand plant
communities with different hydroperiods in the Big Cypress Swamp region of
Florida Everglades. Ecohydrolojjy & Hydro bio lojjy 14: 253-266.

Villa, J. A. and W. J. Mitsch. 2015. Carbon sequestration in different wetland plant
communities in Southwest Florida. International Journal for Biodiversity Science,
Ecosystems Services and Management, doi.org/10. 1080/21 5 13732. 2014. 973909.

Waddington, J. M., and N. T. Roulet. 1997. Groundwater flow and dissolved carbon
movement in a boreal peatiand. Journal of Hydrology 191: 122-138.

Whalen, S. C. 2005. Biogeochemistry of methane exchange between natural wetlands
and the atmosphere. Environmental Engineering Science 22: 73-94.

Whiting, G. J., and J. R Chanton. 2001. Greenhouse carbon balance of wetlands:
Methane emission versus carbon sequestration. Tellus 53B: 521-528.

World Wildlife Fund. 2012. Blue carbon: A new concept for reducing the impacts
of climate change by conserving coastal ecosystems in the Coral Triangle.
WWF-Australia, Brisbane, Queensland, Australia. 20 pp.

Yagi, K., and K. Minami. 1990. Effect of organic matter application on methane
emission from some Japanese paddy fields. Soil Science and Plant Nutrition 36:
599-610.

Yu, K., S. R Faulkner, and M. J. Baldwin. 2008. Effect of hydrological conditions on
nitrous oxide, methane, and carbon dioxide dynamics in a bottomland hardwood
forest and its implication for soil carbon sequestration. Global Change Biology 14:
798-812.

Chapter 18

Wetland Creation and Restoration

Loss rates of wetlands around the world and the subsequent reco£inition of
wetland values have stimulated restoration and creation of these systems.

Policies such as “no net loss” of wetlands in the United States have made
wetland creation and restoration a veritable industry in that country. Wetland
restoration involves returning a wetland to its original or previous wetland
state, whereas wetland creation involves conversion of uplands or shallow
open-water systems to vegetated wetlands. Wetland restoration and creation
can occur for replacement of habitat, for coastal restoration, and for
restoration of mined peatlands. Wetland mitigation banks may overcome many
of the limitations of current approaches to replacing lost wetlands but they are
controversial. Generally, wetland restoration and creation first involve
establishment or reestablishment of appropriate natural hydrologic conditions,
followed by establishment of appropriate vegetation communities. Although
many of these created and restored wetlands have become functional, there have
been some cases of failures of created or restored wetlands, generally caused by a
lack of proper hydrology. Creating and restoring wetlands should be based on
the concept of self-design, whereby any number of native propagules can be
introduced, but the ecosystem adapts and changes itself according to its physical
constraints, and success should not be determined solely by specific plant and
animal presence. Giving these systems sufficient time to carry out their
self-design is another factor that is generally overlooked.

There are two general starting points for anyone interested in getting involved in

wetland restoration and creation:

1 . Learn and understand wetland science and its principles first.

2. Broaden your horizons beyond the field that you were trained in so that you
resist the ever-present temptation to overengineer, overbotanize, or
overzoologize the wedands that you create and restore.

591

592 Chapter 18 Wetland Creation and Restoration

The principles and practices of wetland creation and restoration are based on wet-
land science (hydrology, biogeochemistry, adaptations, and succession). Our advice if
you are interested in creating and restoring wetlands is to first become an expert in
wetland science. Know how the real wetlands work first. That was the intent of the
first 17 chapters in this book. Only after you understand the function and structure of
natural wetlands are you qualified to create and restore wetlands.

The second point is one that needs to be emphasized to all professions. Most
of us have been taught in our lives and professions that we can improve on nature.
Indeed, human civilization is based on that premise. But when we are attempting to
create or re-create natural ecosystems. Mother Nature is in control. In all situations
of wetland creation and restoration, human contribution to the design of weriands
should be kept simple and should strive to stay within the bounds established by the
natural landscape.

The literature on wetland creation and restoration continues to explode, and it is
impossible for us to include all of the possible principles, case studies, and techniques in
this chapter. A critique of the policies and techniques of wetland creation and restora-
tion in the United States was published as a National Academy of Sciences report
(NRC 2001). Mitsch (2013) provided a summary ofwetland creation and restoration
and several case studies from around the world, some of which are updated here. Sev-
eral of these case studies were later reviewed and received “mid-term grades” (Mitsch,
2014). Some notable papers discuss specific wetland creation and restoration projects
for salt marshes (Alphin and Posey, 2000; Craft et ah, 2002; Edwards and Proffitt,
2003; Callaway and Zedler, 2004; Peterson et al., 2005), mangrove swamps (Lewis,
2005; Lewis and Gilmore, 2007), freshwater marshes (Atkinson et ah, 2005; Mitsch
et al., 2012, 2014), peatlands (Gorham and Rochefort, 2003), and forested wetlands
(Rodgers et ah, 2004).

Definitions

Several terms are frequently used in connection with the creation and restora-
tion of wetlands. Precise definitions are important, and confusion about the
exact meaning ofwetland creation, restoration, and related terms is common
(Lewis, 1990a). Bradshaw (1996) concurred that “we must be clear in what
is being discussed."

Wetland restoration refers to the return of a wetland from a disturbed or
altered condition caused by human activity to a previously existing condition.
The wetland may have been degraded or hydro logically altered, and restoration
then may involve reestablishing hydrologic conditions to reestablish previous
vegetation communities.

Wetland creation refers to the conversion of a persistent upland or shallow
water area into a wetland by human activity.

Wetland enhancement refers to a human activity that increases one or
more functions of an existing wetland.

Mitigating Wetland Habitat Loss 593

One type of created wetland, a constructed wetland, refers to a wetland
that has been developed for the primary purpose of contaminant or pollution
removal from wastewater or runoff.

This last type of wetland is also referred to as a treatment wetland and is
the main topic discussed in Chapter 19: “Wetlands and Water Quality."

Significant efforts now focus on the voluntary restoration and creation of wet-
lands. Part of the interest in wetland creation and restoration stems from the fact that
we are losing or have lost so much of this valuable habitat (see Chapter 3: “Wedands
of the World”). Often interest is less voluntary and more in response to government
policies, such as “no net loss” in the United States, that require the replacement of
wetlands for those unavoidably lost. New Zealand, which has lost 90 percent of its
wetlands, has major efforts under way to restore marshes and other wetlands in the
Waikato River Basin on North Island and in the vicinity of Christchurch on South
Island. In southeastern Australia, restoration of the Murray-Darling watersheds, par-
ticularly the riverine billabongs, has become a major undertaking, while coastal plain
wetland restoration and creation are occurring in southwestern Australia.

There are concerted efforts to restore mangrove forests in the Mekong Delta of
Vietnam, along South American coastlines where shrimp farming has destroyed thou-
sands of hectares of mangroves, and around the Indian Ocean to provide tsunami and
typhoon protection for coastal areas. Tidal marshes have been created along much of
China’s eastern coastline, and wedand creation and restoration are now occurring in
the Yangtze Delta in Shanghai and upstream of the Three Gorges Dam on the upper
Yangtze. Wetland restoration and creation are being proposed or implemented on very
large scales to prevent more deterioration of existing wetlands (Everglades in Florida),
to mitigate the loss of fisheries (Delaware Bay in eastern United States), to reduce land
loss and provide protection from hurricanes (Mississippi Delta in Louisiana), to stabi-
lize a watershed and provide water quality improvement (Skjern River, Denmark), and
to solve serious cases of overenrichment of coastal waters (Baltic Sea in Scandinavia;
Gulf of Mexico in United States; Laurentian Great Lakes).

Mitigating Wetland Habitat Loss

Wetland protection regulations in the United States and now elsewhere have led to
the practice of requiring that wetlands be created, restored, or enhanced to replace
wetlands lost in developments such as highway construction, coastal drainage and fill-
ing, or commercial development. This is referred to as the process of “mitigating” the
original loss, and these “new” wetlands are often called mitigation wetlands. (Note:
To mitigate means to “make less harsh or harmful.” The term mitigation wetland or
wetland mitigation is therefore poor use of English. We should rather refer to “miti-
gating the loss of a wetland.”) Perhaps it might be more appropriate to refer to these
wetlands as replacement wetlands.

594 Chapter 18 Wetland Creation and Restoration

Figure 18.1 Proper wetland mitigation with comparisons with both what has been lost (legal
success) and with regional reference natural wetlands (ecological success). (From Wilson
and Mitsch, 1996)

Figure 18.1 illustrates conceptually how success should be measured for replace-
ment wetlands. Le^al success involves a comparison of the lost wetland function and
area with that which is gained in the replacement wetland. Ecological success should
involve a comparison of the replacement wetland with a reference wetland (natural
wetlands of the same type that may occur in the same setting or generally accepted
“standards” of regional wetland function). Overall success would then be gauged by
a combination of the legal and ecological comparisons. While this model represents
an ideal, the comparison involving both standards is rarely done.

In fact, the usual decision is based on the size of the wetland lost and little
else. Replacement wetlands are designed to be at least the same size as the lost wet-
lands, but more often a mitigation ratio is applied so that more wetlands are created
and/or restored than are lost. For example, a mitigation ratio of 2:1 means that
2 hectares (ha) of wetlands will be restored or created for every hectare of wetland lost
to development. Considerable controversy exists, for example, in the United States,

Mitigating Wetland Habitat Loss 595

on the question as to whether wetland loss can be mitigated successfully or if it is
essentially impossible (NRC, 2001). Robb (2002) reviewed several years’ efforts on
mitigating wetland loss in Indiana and suggested, based on failure rates of various wet-
land types, that there should be these mitigation ratios: 7.6:1 for wet meadows, 3.5:1
for forested wetlands, 1.2:1 for freshwater marshes, and 1:1 for open-water systems.

On paper, the U.S. Army Corps of Engineers’ implementation of the U.S. policy
of “no net loss” of wetlands over the past 20 years (see Chapter 15: “Wetland Laws and
Protection”) appears to be working. There was an estimated net gain of 8,000 ha/yr of
wetlands and associated uplands in the United States over the 20-year period of 1993
through 2012 as a result of enforcement of the Clean Water Act through mitigation
of wetland loss (Fig. 18.2). This number is the result of the issuing of permits for the
destruction of 8,000 ha/yr of wetlands and the creation, restoration, enhancement,
or preservation of approximately 16,000 ha/yr of wetlands and associated uplands.
Over those 20 years of record, the United States has lost 161,000 ha of wetlands and
“gained” 318,000 ha of mitigation credit.

There are two reasons why one should not be so euphoric about this account-
ing that shows a net gain of wetlands. First, it is impossible to tell from these general
numbers just how successful this wetland trading has been, because few statistics exist
on what functions were lost versus what functions were gained. There is some dis-
cussion of this at the end of this chapter. Second, the estimated gain of 157,000 ha
over 20 years does not make much of an impact on the loss of 47,000,000 ha of
wetlands that occurred from presettlement time up to the 1980s in the United States.
There also appears to a distinct pattern of fewer wetlands permitted for drainage and
therefore less mitigation in recent times (Fig. 18.2). About 11,000 ha/yr of wetlands
were created or restored from 2007 to 2012 compared to 20,000 ha/yr from 1996 to

25000

20000

15000

S 10000

5000

□ permitted
■ mitigated

^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^ ^

Figure 18.2 Patterns of mitigation for wetiand loss in the United States for the 20-year
peried 1993 to 2012. “Permitted” refers to the area of wetiands that were permitted for
drainage in a given year. “Mitigated” refers to the area of wetiands that were required as
compensatory mitigation. (Data courtesy ef Pubiic Affairs Office, U.S. Army Corps of Engi-
neers, Washington, DC)

596 Chapter 18 Wetland Creation and Restoration

2005. The lower numbers could be because of less wetland protection at the federal
level as a result of the 2001 and 2006 U.S. Supreme Court decisions described in
Chapter 15. Some of that decrease can also be explained by economic downturn in
the USA and hence fewer housing and other developments in the late 2000s.

Mitigation Banks

One of the more interesting strategies that the private sector and government agencies
have developed to deal with the piecemeal approach to mitigation of wedand loss is
the concept of a mitigation bank. A mitigation bank is defined as “a wetland area that
has been restored and protected to provide compensation for impacts to wedands”
(USEPA and US Army Corps of Engineers, 2008). In this approach, wetlands are
usually built in advance of development activides that cause wedand loss, and credits
of wedand area can be sold to those who are in need of midgation for wedand loss.
Banks are seen as a way of streamlining the process of midgadng wedand loss and, in
many cases, providing a large, fully funcdonal wedand rather than small, quesdonable
wedands near the site of wedand loss. The midgadon bank can be set up with bonds
ensuring compliance. Arrangements are easier for wetland mitigation banks to be man-
aged “in perpetuity” through conservadon easements or transfer of titles to resource
agencies. Financial resources can be arranged ahead of time for proper monitoring
of the wedand bank. Midgadon banks can be publicly or privately owned, although
there is a potendal conflict of interest if public agencies run midgadon banks. Public
agencies could be involved in enforcing reguladons on midgadng wedand loss and
then steer permittees toward their own banks rather than to private banks.

In 1992, there were only 46 wedand midgation banks in the United States. By
2002, there were 219 mitigation banks, both private and public, covering 50,000
ha in 29 states in the country (Spieles, 2005). As of August 2013, there were over
1,800 mitigation bank sites listed in the United States. At one time, there were 62
formal midgadon banks (proposed and operadng) and hundreds of quasi-mitigadon
banks in Florida alone (Ann Redmond, personal communication). It appears that,
if midgadon of wedand loss condnues to be the nadon’s policy and if regulation of
mitigadon banks can be developed that is fair and uncomplicated, the use of midgadon
banks to solve this “wedand trading” issue will condnue to increase well through the
twenty-first century.

New regulations on midgadng the loss of wedands were developed in 2008 by
the U.S. Army Corps of Engineers and U.S. EPA (2008). These reguladons were
meant to increase the effecdveness of wedand mitigadon banks and to strengthen
requirements for the use of in -lieu fee midgadon (U.S. EPA and U.S. Army Corps of
Engineers, 2008). The data in Figure 18.2 do not show any obvious improvement in
the midgadon ratios after 2008 yet.

Agricultural Land Restoration

For many decades in the United States, farm pond creadon was encouraged as a way
of providing drinking water for domesdc animals and other funcdons on the farm.

Forested Wetland Restoration 597

Although individually quite small (usually about 0.2 ha), the total number of con-
structed ponds is large. Several years ago, ponds were being constructed at a rate of
about 50,000 per year. Marshes often develop around the perimeter of many of these
ponds, while other ponds have converted to marshes. Many of these ponds were built
with large, shallow areas to attract waterfowl, and these shallow zones have become
typical pothole marshes.

Dahl (2006) estimated that between 1998 and 2004, wetland pond areas
increased by 280,000 ha in the United States, a 12.6 percent increase. Most of
this gain (141,000 ha) resulted on nonagricultural upland, while 29,000 ha were
constructed on farmland. Many of the nonagricultural ponds are built in housing
and commercial developments, especially in states like Florida, as stormwater runoff
ponds. There are those who question the ecological value of these ponds; for
example, some regulators do not like ponds because they have fish and therefore
cannot support amphibians.

Conservation programs were set up in the 1990s to encourage individual farm-
ers in the United States to restore wedands on their land. Both the Conservation
Reserve Program (CRP) and the Wetlands Reserve Program (WRP) under the U.S.
Department of Agriculture (USDA) have led to significant areas of wetlands being
restored or protected. CRP guidelines, announced in 1997, give increased emphasis
to the enrollment and restoration of cropped wetlands — that is, wetlands that produce
crops but serve wetland functions when crops are not being grown. The CRP also
encourages wetland restoration, particularly through hydrologic restoration.

The WRP, a voluntary program established in 1992, and was specific for wetland
restoration; it offered landowners the opportunity to protect, restore, and enhance
wetlands on their property and provides funds for farmers to do so. The USDA Nat-
ural Resources Conservation Service provides technical and financial support to help
landowners. The WRP options to protect, restore, and enhance wetlands and associ-
ated uplands include permanent easements, 30-year easements, or 10-year restoration
cost-share agreements. As of 2012, approximately 900,000 ha of wetlands and adja-
cent uplands have been enrolled in the WRP in the United States, with the most
intense activity in the lower Mississippi River basin and Florida. The U.S. Agricultural
Act of 2014 established the Agricultural Conservation Easement Program (ACEP) in
the U.S. Department of Agriculture. This act repealed the WRP but did not affect
the validity or terms of any WRP agreement entered into prior to February 7, 2014.
Wetland reserve easements to restore, protect, and enhance wetlands continue under
the ACEP. The long-range affect that this administrative change will have on wetland
conservation and restoration is unknown.

Forested Wetland Restoration

There is less experience with forested wetland restoration and creation compared to
herbaceous marshes, although these wetlands have been lost at alarming rates, par-
ticularly in the southeastern United States. Forested wetland creation and restoration
are different from marsh creation and restoration, because forest regeneration takes
decades rather than years to complete, and there is more uncertainty about the results.

598 Chapter 18 Wetland Creation and Restoration

Much riparian forest restoration in the United States has centered on the 10-million-ha
lower Mississippi River alluvial valley, where more than 182,000 to 220,000 ha have
been reforested (Haynes, 2004), primarily with bottomland hardwood species and, to
a lesser extent, deepwater swamp species. This is a small contribution to the restora-
tion of this alluvial floodplain, where 7.2 million ha of bottomland hardwood forest
were estimated to have been lost (Hefner and Brown, 1985).

Hydrologic Restoration of Wetlands

Lines often blur between wetlands created and restored for habitat restoration and
those restored for water quality and hydrology improvement. In fact, most wedands
that are restored or created are done so for both reasons. One of the largest freshwater
wetland restorations in the world is being carried out in the Florida Everglades to
restore, at least to some degree, the natural hydrologic conditions, at least in the
Everglades that are left (see Case Study 1). The restoration of the Mesopotamian
Marshlands of Iraq (Case Study 2) is an example where hydrology was restored to
a drained wetland of incredible cultural importance. There, the hydrology had been
purposefully disrupted by the regime of Saddam Hussein and indirectly through
upstream river management by Iraq’s upstream neighbors. The Iraqi people, with
some international assistance, are undertaking a hydrologic restoration of this histori-
cally and culturally important wetland (see also Chapter I : “Wetlands: Human Use and
Science”).

CASE STUDY 1: Restoring the Florida Everglades

The restoration of the Florida Everglades, one of the largest wetland areas in
the world, actually involves several separate initiatives being carried out in the
4.6-million-ha Kissimmee-Okeechobee-Everglades (KOE) region in the south-
ern third of Florida (Fig. 18.3). The basic plan involves restoring something
closer to the original hydrology of the KOE region (Fig 18.3a, b) by sending
less of the water from the upper watershed to the Caloosahatchee River to
the west and the St. Lucie Canal to the east than is currently the case and
directing more of the water to the Everglades south of Lake Okeechobee.

The actual hydrologic flows for water years 2012 and 2013 (Fig. 18.3c
and Table 18.1) were not close to either the historic or restoration goals of
Figure 18.3a, b. Despite the fact that precipitation was slightly below normal
in 2012 and almost exactly at normal amounts in 2013 (1,350 mm/yr), a
significant portion of the Lake Okeechobee outflows were still shunted down
the Caloosahatchee to the west and the St. Lucie Canal to the east in 2013:

Hydrologic Restoration of Wetlands 599

746,000 m^yr, or 58 percent of the 1,285,000 mVyr that was diverted south
into the Florida Everglades. There has been significant pollution and ecological
problems in the estuaries of the Gulf of Mexico and Atlantic Ocean because
of these excessive freshwater flows from Lake Okeechobee going east and
west instead of south in the Florida Everglades as proposed by the original
restoration goal shown in Figure 18.2c.

Figure 18.3 Illustrations of Florida Everglades wetland restoration: (a) historic con-
ditions of the Florida Everglades hydrology in presettlement times; (b) the flow con-
ditions desired when the Everglades is restored; (c) flow conditions in water years
2012 and 2013 in the Everglades where much of the water was still sent east and
west to the sea rather than south to the Everglades (flows for numbered pathways are
given in Table 18.1); (d) captured Burmese python {Python molurus bivittatus) in Florida
Everglades as symbolic of major problems of invasive species that continue to occur,
((a), (b) from Mitsch and Jorgensen, 2004; (c) from South Florida Water Management
District (SFWMD, 2014); (d) from Mike Rochford, reprinted with permission)

600 Chapter 18 Wetland Creation and Restoration

Figure 18.3 {Continued)

Table 18.1 Water flow through the Florida Everglades for water years 2012 and
2013 as illustrated in Figure 18.3c (Flows are x 1000 m^/yr.) Flow pathway number
refers to pathways shown in Figure 18.3c

Flow pathway in Greater Florida Everglades

Water Year 2013 Water Year 2012

1. Lake Kissimmee outflows

543

1, 004

2. Lake Istokpoga outflows

347

281

3. Lake Okeechobee Inflows

2,590

2,246

4. Lake Okeechobee outflows

1,285

920

5. Flows into the St. Lucie Canal from Lake Okeechobee

128

6. Flows into the St. Lucie Estuary through the St. Lucie Canal

189

7. Flows into the Caloosahatchee Canal from Lake Okeechobee

618

222

8. Flows into the Caloosahatchee Estuary through the
Caloosahatchee Canal

1,404

739

9. Water Conservation Area 1 inflows

449

210

10. Water Conservation Area 1 outflows

597

11. Water Conservation Area 2 inflows

1,325

476

12. Water Conservation Area 2 outflows

1,151

466

13. Water Conservation Area 3 inflows

1,631

1,110

14. Water Conservation Area 3 outflows

1,511

704

15. Everglades National Park inflows

1,847

918

Source: SFWMD (2014).

Hydrologic Restoration of Wetlands 601

Specific problems in the Everglades have developed for four reasons:

1. Excessive nutrient loading to Lake Okeechobee and to the
Everglades, primarily from agricultural runoff

2. Loss and fragmentation of habitat caused by urban and agricultural
development

3. Spread of cattail (Typha), melaleuca {Melaleuca quinquenervia), and
other invasives and exotics in the Everglades, replacing native
vegetation

4. Hydrologic alteration due to an extensive canal and straightened river
system built by the U.S. Army Corps of Engineers and others for flood
protection, and maintained by water management districts

One major restoration project in the KOE region that initially received a
lot of attention is the restoration of the Kissimmee River. As a result of the
channelization of the river in the 1960s, a 166-km-long river was transformed
into a 90-km-long, 100-m-wide ditch, and the extent of wetlands along the
river decreased by 65 percent (Table 18.2). Waterfowl populations decreased
by 90 percent as a result of this channelization (Blake, 1980). The restoration
of the Kissimmee River is a major undertaking to reintroduce the sinuosity
to the artificially straightened river. The river restoration work, expected to
be completed in stages by 2019, will return 50 km^ of lost wetland habitat
to the riparian zone and reestablish 64 km of the river, at a total expected
cost of almost US$1 billion (Koebel and Bousquin, 2014). It will also provide
sinks for nutrients that are otherwise causing increased eutrophication in
downstream Lake Okeechobee. Restoration to date is described in a special
issue of Restoration Ecology (Bousquin, 2014). NRC (2014) concluded
that the Kissimmee River restoration “is probably the most advanced
in demonstrating substantial restoration of the natural system” in the

Table 18.2 Wetland changes due to channelization of the Kissimmee River in
south Florida (Channelization took place between 1962 and 1971 and
transformed a 166-km meandering river into a 90-km-long, 10-m-deep,
100-m-wide canal.)

Wetland Type

Prechannelization (ha)

Postchannelization (ha)

Percentage Change (%)

Marsh

8, 892

1, 238

-86

Wet prairie

4, 126

2,128

-48

Scrub-shrub wetland

2, 068

1,003

-51

Forested wetland

150

243

+62

Other

533

919

+72

Total

15, 769

5,531

-65

Source: Toth et al. (1995).

602 Chapter 18 Wetland Creation and Restoration

Florida Everglades and that its long-term monitoring of restoration progress
is a useful example for other restoration projects in the Florida Everglades.

Everglades restoration also involves halting the spread of high-nutrient
cattail {Typha domingensis) through the low-nutrient sawgrass (Cladium
jamaicense) communities that currently dominate the Everglades (see
Chapter 14: “Human Impacts and Management of Wetlands” for general
description of the water pollution problem in the Everglades and Chapter 19:
“Wetlands and Water Quality” for a description of the created and restored
wetlands being used to solve this problem).

Overall, the Everglades restoration, as now planned by the U.S. Army
Corps of Engineers, has a budget of over $20 billion and may take decades
more to complete.

A not-unrelated aspect of the restoration of the Florida Everglades is deal-
ing with the vast number and often dominance of invasive species that have
become entrenched there over the past century. There are about 250 nonna-
tive plants in the ecosystems of the Florida Everglades, or about 16 percent
of the total species count. Of these, 12 species are considered particularly
important to the Everglades restoration (Rodgers et al., 2014). This short list
includes plants that were purposefully introduced to south Florida and are now
the targets of continual removal attempts in and around the Everglades, such
as melaleuca, introduced in 1906 from Australia to enhance drainage of the
Everglades, and Brazilian pepper {Schinus terebinthifolius), which was brought
to Florida from South America as an ornamental in the mid-nineteenth century.

The invasive animal that has received more news media attention than
any other is the Burmese python (Python molurus bivittatus) (Fig. 18.3d). The
snakes, which can achieve lengths of up to 5.5 m, have been suggested to
have a population as large as several thousand individuals in south Florida
(M. Dorcas, personal communication, 2014). Dorcas et al. (2012) illustrated
that over a 90 percent reduction in populations or mammals such as raccoons,
opossums, bobcats, and rabbits occurred in the Everglades National Park coin-
cident with the significant increase in python populations since 2000. Partly
as a result of this study and the follow-up media attention given to this subject,
a “Python Challenge” was sponsored in the winter 2013 in south Florida. Hun-
dreds of hunters from all over the United States and Canada spent a month
trying to capture and/or kill the reptile. In the end, only 68 snakes were killed
or captured (http://phys.org/news/2013-02-python-everglades-nets.html).

NRC (2014) assessed the progress of the Florida Everglades restoration
and discussed specific scientific and engineering issues that may impact
further progress. They recommended a dedicated source of funding that
could provide “ongoing long-term system-wide monitoring and assessment”
of progress on restoration objectives. The report makes recommendations
for restoration activities, project management strategies, management of

Hydrologic Restoration of Wetlands 603

invasive nonnative species, and high-priority research needs. Among the
recommendations are “finding solutions to overcome current constraints
related to authorization, funding, and water quality permitting” and dealing
in a more programmatic way with climate change and sea-level rise that the
report suggests “provide even more incentive for restoring the Everglades
ecosystem." The report also brings attention to the importance of dealing with
invasive species in the Everglades, a problem that did not exist prior to the
drainage projects of the last century. The report concludes that “restoration
progress made by CERP [Comprehensive Everglades Wetland Plan] projects
to date remains fairly modest in scope.” In fact, the authors note that some
of the most Important progress was with the Kissimmee River restoration
and the Clll South Dade County spreader canal project, neither of which is
part of the CERP

CASE STUDY 2: Restoration of the Mesopotamian Marshiands

The Mesopotamian Marshlands of southern Iraq and Iran were described in
Chapters 1 and 3. These wetlands, found at the confluence of the historic
Tigris and Euphrates rivers, were 15,000 to 20,000 km^ in area as recently
as the early 1970s (Fig. 18.4a) but were drained and diked, especially in the
1990s, to less than 10 percent of that extent by 2000 (Fig. 18.4b). Among
the main causes are upstream dams and drainage systems constructed in the
1980s and 1990s that altered the river flows and eliminated the flood pulses
that sustained the wetlands.

Since the overthrow of Saddam Hussein’s dictatorship in Iraq in 2003,
there has been a concerted effort by the Iraqis and then the international com-
munity at restoring the marshlands (Richardson et al., 2005). The restoration
has often occurred with local residents breaking dikes or removing impedi-
ments to flooding. Remote-sensing images showed that at least 37 percent of
the wetlands were restored by 2005. Frontiers in Ecology and the Environment
(3, No. 8, October 2005) reported In 2005 that “at least 74 species of migra-
tory waterfowl and many endemic birds have been sighted in a survey of Iraq’s
marshland.” It was also reported that as many as 90,000 Marsh Arabs have
returned to the wetlands already (Azzam Alwash, personal communication).

Alwash, the director of the Eden Again effort, has suggested that perhaps
as much as 75 percent of the marshlands can be restored (Fig. 18.4c, d).
Several questions still remain unanswered about whether full restoration can
occur, including whether adequate water supplies exist in the rivers, given the
competition from Turkey, Syria, and Iran, and within Iraq itself, and whether

604 Chapter 18 Wetland Creation and Restoration

landscape connectivity of the marshes can be reestablished (Richardson and
Hussain, 2006).

Figure 18.4 The Mesopotamian Marshiands of iraq, with shading indicating extent of
the marshlands: (a) in 1970 before extensive drainage; (b) in 2000 after extensive
drainage; (c) as expected in the future with 75 percent restoration of the marshland;
and (d) photo of the Mesopotamian Marshland after restoration, ((a) (b) from UNEP,
2001; (c), (d) from Azzam Alwash)

Peatland Restoration 605

Peatland Restoration

Peatland restoration is a relatively new type of wetland restoration compared to other
types and potentially could be the most difficult (Gorham and Rochefort, 2003).
Early attempts with peatlands occurred in Europe, specifically in Finland, Germany,
the United Kingdom, and the Netherlands. Increased peat mining in Ganada and else-
where has led to increased interest in understanding if and how mined peatlands can be
restored. When peat surface mines are abandoned without restoration, the area rarely
returns through secondary succession to the original moss -dominated system (Quinty
and Rochefort, 1997). There is some promise that peadand restoration can be success-
ful (Rochefort and Lode, 2006), but because surface mining causes major changes in
local hydrology and peat accumulates at an exceedingly slow rate, restoration progress
will be measured in decades rather than years.

In the 1960s and 1970s, block harvesting of peat was replaced by vacuum har-
vesting in southern Quebec and in New Brunswick, necessitating the development
of different restoration techniques. While traditional block-cutting of peat left a vari-
able landscape of high ground and trenches, vacuum harvesting leaves relatively flat
surfaces bordered by drainage ditches. Abandoned block-cut sites appear to revege-
tate with peatland species more easily than do vacuum-harvested sites, and the latter
can remain bare for a decade or more after mining (Rochefort and Gampeau, 1997).
Gase Study 3 describes as peadand restoration research field lab in Quebec where the
best ways to restore peatlands are being investigated.

CASE STUDY 3: Peatland Restoration Research at Bois-des-Bel,
Quebec

Despite the vast expanses of peatlands in the world, whole-ecosystem exper-
iments on restoration research on this type of wetland are rare. Bois-des-
Bel peatland, located about 200 km northeast of Quebec City, on the
southern shore of the St. Lawrence River in Quebec, Canada (Fig. 18.5),
is a whole-ecosystem research site where scientists are evaluating the
pace of peatland restoration after peat mining (Rochefort et al., 2003;
Waddington et al., 2008). The entire peatland is about 210 ha; the research
peatland area is about 11.5 ha that was drained in 1972 and mined by a
vacuum extraction technique from 1973 to 1980. When mining stopped, a
2-m peat deposit remained. Restoration began in 1999 on 8.4 ha of the
site, with the remaining as an unrestored control. The restored area was
divided into four zones, each of which has two shallow pools (13m x 5m
X 1.5 m max depth) for aquatic and amphibian habitat. Line Rochefort and
her students and colleagues at Universite Laval, Quebec City, and at other
Canadian universities have established the site as a long-term ecosystem
research site to investigate the revegetation of mined peatlands (Price
et al., 1998; Rochefort et al., 2003, 2012; Waddington et al., 2003, 2008;

606 Chapter 18 Wetland Creation and Restoration

Isselin-Nondedeu et al., 2007). The restoration involved terracing to produce
better water distribution, reintroduction of Sphagnum diaspores harvested
from a nearby natural wetland, and reflooding by blocking drainage ditches.
The moss carpet increased by about 12 cm by 2007 and was three times the

(a)

Figure 18.5 Bois-des-Bel experimental peatland on the southern shore of the St.
Lawrence River in Quebec, Canada, (a) map showing restored (west) and cutover
(east) sites including four research zones in restored area for vegetation monitoring
in created pool at Bois-des-Bel restored site; (b) shallow pools created in the restored
area for aquatic and amphibian habitat, ((a) from Waddington et al., 2008; (b) photo
courtesy of L. Rochefort, Peatland Ecology Research Group (PERG), Universite Laval,
Quebec City, Canada, reprinted by permission)

Coastal Wetland Restoration 607

(b)

Figure 18.5 {Continued)

thickness that it was in 2003. Sphagnum cover by 2005 was 60 percent of the
area in the restored sites compared to oniy 0.25 percent in the nonrestored
sites (isseiin-Nondedeu et ai., 2007). The restored sites aiso exported iess
than haif the dissoived organic carbon than did the cutover peatiand sites
(Waddington et ai., 2008). This area has been monitored postrestoration for
aimost a decade (Rochefort et ai., 2013).

Coastal Wetland Restoration

Salt Marsh Restoration

There is a great deal of interest in coasdine restoration. Early pioneering work
on salt marsh restoration was done in Europe (Lambert, 1964; Ranwell, 1967);
China (Chung, 1982, 1989); and the United States on the North Carolina coastline
(Woodhouse, 1979; Broome et al., 1988), in the Chesapeake Bay area (Garbisch
et al., 1975; Garbisch, 1977, 2005), and along the coastlines of Florida, Puerto Rico
(Lewis, 1990b, 1990c), and Galifornia (Zedler, 1988, 2000b; Josselyn et al., 1990).
Some of this coastal wetland restoration has been undertaken for habitat development
as mitigation for coastal development projects.

For coastal salt marshes in the eastern United States, the cordgrass Spartina
alterniflora, is the primary choice for coastal marsh restoration, but the same species
is considered an invasive and unwanted plant on the West Goast of North America.
Spartina townsendii, S. anpilica-, and S. alterniflora have been used to restore salt
marshes in Europe, New Zealand, and Ghina, although several of these species are
now considered invasive in those regions. Salt marsh grasses tend to distribute easily
through seed dispersal, and the spread of these grasses can be quite rapid once the
reintroduction has begun, as long as the area being revegetated is intertidal — that is.

608 Chapter 18 Wetland Creation and Restoration

the elevation is between ordinary high tide and low tide. The details of successful
coastal wedand creation are site specific, but six generalizations seem to be valid in
most situations:

1 . Sediment elevation is the most critical factor determining the successful
establishment of vegetation and the plant species that will survive. The site
must be intertidal.

2. In general, the upper half of the intertidal zone is more rapidly vegetated
than lower elevations.

3 . Sediment composition does not seem to be a critical factor in colonization by
plants unless the deposits are almost pure sand that is subject to rapid
desiccation at the upper elevations.

4. The site needs to be protected from high wave energy. It is difficult or
impossible to establish vegetation at high-energy sites.

5. Most sites revegetate naturally from seeds if the elevation is appropriate and
the wave energy is moderate. Sprigging live plants has been accomplished
successfully in some cases, and seeding also has been successful in the upper
half of the intertidal zone.

6. Good stands can be established during the first season of growth, although
sediment stabilization does not occur until after two seasons. Within tour
years, successfully planted sites are often indistinguishable superficially from
natural marshes.

Several early studies emphasized the importance of restoring tidal conditions,
including salinity, to marsh areas that had become more “freshwater” because of iso-
lation from the sea. In cases such as this, the restoration is simple: Remove whatever
impediment is blocking tidal exchange. Case Study 4 describes a salt marsh restora-
tion where this has been done with some success: When the natural tidal hydrology
was restored, the vegetation and aquatic species followed. Case Study 5 describes a
many decade restoration of a coastal river/wedand complex in the New York City
urban region.

CASE STUDY 4: Delaware Bay Salt Marsh Restoration

A large coastal wetland restoration project in the eastern United States
involves the restoration, enhancement, and preservation of 5,000 ha
of coastal salt marshes on Delaware Bay in New Jersey and Delaware
(Figure 18.6a). This estuary enhancement, being carried out by New Jersey's
electric utility (Public Service Enterprise Group [PSEG]), with advice from a
team of scientists and consultants, was undertaken as mitigation for the

Coastal Wetland Restoration 609

potential impacts of once-through cooling from a nuclear power plant operated
by PSEG on the bay. The reasoning was that the impact of once-through
cooling on fin fish, through entrainment and impingement, could be offset
by increased fisheries production from restored salt marshes. Because of
uncertainties involved in this kind of ecological trading, the area of restoration
was estimated as the salt marshes that would be necessary to compensate
for the impacts of the power plant on fin fish times a safety factor of 4. Three
distinct approaches are being utilized in this project to restore the Delaware
Bay coastline:

1. Reintroduce flooding. The most important type of restoration involves
the reintroduction of tidal inundation to about 1,800 ha of former
diked salt-hay farms. Many marshes along Delaware Bay have been
isolated by dikes from the bay, sometimes for centuries, and put into
the commercial production of “salt hay” {Spartina patens). Hydrologic
restoration was accomplished by excavating breaches in the dikes
and, in most cases, connecting these new inlets to a system of
re-created tidal creeks and existing canal systems.

2. Reexcavate tidal marshes. Additional restoration involves enhancing
drainage by reexcavating higher-order tidal creeks in these newly
flooded salt marshes, thereby increasing tidal circulation. This is
particularly important in marshes that were formerly diked, because
the isolation from the sea has led to the filling of former tidal creeks.
After initial tidal creeks were established, it was expected that the
system would self-design more tidal channels and increase the
channel density.

3. Reduce Phragmites domination. In another set of restoration sites in
Delaware and New Jersey, restoration involves the reduction in cover
of the aggressive and invasive reed grass (Phragmites australis) in
2,100 ha of nonimpounded coastal wetlands. Alternatives that were
investigated include hydrological modifications such as channel
excavation, breaching remnant dikes, microtopographic changes,
mowing, planting, and herbicide application.

Results of this study were reported in several presentations and reports
and several early Journal articles, including Teal and Weinstein (2002), and
several papers in a special issue of Ecological Engineering (Peterson et al.,
2005). From a hydrodynamic perspective, in those marshes where tidal
exchange was restored, the development of an intricate tidal creek density
from the originally constructed tidal creeks has been impressive. Figure 18.6b
illustrates the development of a stream network at one of the newly restored

Figure 18.6 Delaware Bay salt marsh restoration from 1995 through present: (a) map
of Delaware Bay between New Jersey and Delaware, showing locations of 5,800-ha
salt marsh restoration that is being carried out to mitigate the loss of fin fish due
to entrainment and impingement caused by once-through cooling at a nuclear power
plant. Wetlands are being preserved, restored from salt-hay farms by reintroducing
flooding, and enhanced by removal of Phragmites australis, (b) Total number of stream
channels by channel class at Dennis Township (1996-2014) and Commercial Town-
ship (1997-2013) restored salt marshes on Delaware Bay. (c) Vegetative cover in 1995
prior to restoration at the Dennis Township restored salt marshes. White area indicates
unvegetated region, (d) Vegetative cover in 2003, seven years after low-channel-class
tidal creeks were restored and dikes were breached. Predominant gray areas are
restored Spartina alterniflora and other desirable marsh vegetation, (e) Project con-
sultant John Teal viewing a marsh area mostly restored to Spartina aiternifiora of
a portion of a 640-ha Alloway Creek, New Jersey, marsh on Delaware Bay that was
restored from Phragmites austraiis with herbicide (mostly glyphosate) application over
a period of more than 10 years, ((b) provided with permission, Kenneth A. Strait, PSEG
Service Corporation, Salem, NJ; (c), (d) revised from Hinkle and Mitsch, 2005; (e) photo
by W. J. Mitsch)

610

(b)

Figure 18.6 {Continued)

611

612 Chapter 18 Wetland Creation and Restoration

Vegetative cover categories
■i Sjpa/tina/other desirabte mars^ vegetation
Salt hay field

Phiagmifes dominated vegetation
Deed Phragmites australts

2003

1995

Feet 0 600 1200 )800

i I ‘l I ‘I ! ' I
Met«s 0 200 400 600

I Non-vegeitaled marsh ptan
I Ponded water
I Chanr>el

I Upland/developed land
] Wetland restoration area boundary

(d)

(e)

Figure 18.6 {Continued)

marsh sites — Dennis Township. The “order” of the stream channels increased
from 5 or less to well over 20 from 1996 through 2004. The number of small
tributaries increased from “dozens” to “hundreds” at all three salt-hay farm
sites that were reopened to tidal flushing. For the first three years, there was
a rapid increase in the growth of channel orders 3 through 9; in the next
three years, there was a rapid increase in the channel orders 10 through 16.
(Note: This definition uses channel order as opposite to the normal method on
stream order; here the largest channels are designated as channel order 1.)
Hydrologic design did occur in a self-design fashion after only initial cuts by
construction of the first-order channels.

Coastal Wetland Restoration 613

For the salt-hay farms that were flooded, typical goals include a high per-
centage cover of desirable vegetation such as Spartina alterniflora, a relatively
low percentage of open water, and the absence of the invasive reed grass
Phragmites australis. The success of this coastal restoration project, subject
to a combination of legal, hydrologic, and ecological constraints, is also being
estimated through comparison of restored sites to natural reference marshes.
Results of this part of the project after almost two decades are encouraging.
At the formerly diked salt-hay farms, reestablishment of Spartina alterniflora
and other favorable vegetation has been rapid and extensive.

In Dennis Township, approximately 70 percent of the site was dominated
by Spartina alterniflora after only two growing seasons and almost 80 percent
by the fifth year after construction (Fig. 18.6c, d). Tidal restoration was com-
pleted at the Maurice River site, which is twice the size of the Dennis Township
site in early 1998. Major revegetation by Spartina alterniflora and some Sal-
icornia has already occurred, with 71 percent of the site showing desirable
vegetation after four growing seasons. At the third and the largest salt-hay
farm restoration site, at Commercial Township, which is five times larger than
Dennis Township site, revegetation is occurring rapidly from the bayside. This
study has shown that the speed with which salt marsh restoration takes place
is dependent on three main factors:

1. The degree to which the tidal “circulatory system” works its way
through the marsh

2. The size of the site being restored

3. The initial presence of Spartina and other desirable species

No planting was necessary on these sites, as Spartina seeds arrive by
tidal fluxes, but the design of the sites to allow that tidal connectivity (and
hence the importance of appropriate site elevations relative to tides) was crit-
ical. Self-design works when the proper conditions for propagule disbursement
are provided. Extensive ponding in some areas of the marshes, especially
at Commercial Township, which has the highest ratio of area to edge, has
impeded the reestablishment of Spartina in some locations (Teal and Wein-
stein, 2002). Creating additional streams or waiting for the tidal forces to
cause the same effect eventually allows these areas to develop tidal cycles
and Spartina to establish itself.

Reducing Phragm/tes domination in another set of brackish marsh restora-
tion sites in Delaware and New Jersey has required more years of effort,
mainly through the use of herbicides, but there are now substantial areas
of significant recovery of the marsh to desirable vegetation, including Spartina
alterniflora. The Phragmites cover in the marsh shown in the photo at the
Alloway Creek watershed in New Jersey (Fig. 18. 6e) decreased from 60 per-
cent cover in 1996 to less than 5 percent cover in 2013.

614 Chapter 18 Wetland Creation and Restoration

CASE STUDY 5: Urban Coastal Restoration in the New York City
Region

The Hackensack Meadowlands (Fig. 18.7a) is adjacent to heavily industrial-
ized and commercial Newark Bay, New Jersey, and Is on one of the most
polluted water courses in the United States. It is located a few km west of the

Central

Brooklyn

Borough Parks

1 . Harrier Meadow

2. PollomaQid Mibgabon

3. Mill Creek Marsh

4. Western Brackish Marsh

5. Eastern Brackt^ Marsh

6. Secaucus High School WetlarxJ
Enhancement Site

7. Marsh Resources MeadowtarxJs
Mitigation Bank Phase i & II

6. Evergreen MRi Phase III

9. Kane Natural Area Mitigation Bank

10. Vince Lombardi Marsh

1 1 . Skaetkill Creak Marsh

Newark

I restored wetland sites
\ district boundary

(a)

Figure 18.7 Hackensack Meadowlands, New Jersey, in New York City metropolitan
area: (a) Map of Hackensack Meadowland District in New Jersey and location of wet-
land restoration projects in the district; (b) photo of Hackensack Meadowlands with
New York City skyline in the distance, ((a) Ftom maps provided by New Jersey Meadow-
lands Commission (NJMC)/Meadowlands Environmental Research Institute; (b) photo
by W. J. Mitsch)

Coastal Wetland Restoration 615

Figure 18.7 (Continued)

Hudson River and Manhattan Island in New York City (Fig. 18.7b). More than
a century of ditching, filling, drainage, and diking have changed the lower
reaches of the Hackensack River from a tidal salt and brackish marshland
into highly urbanized mix of residential and industrial land uses interspersed
with tidal creeks and marshes and mudflats. The Hackensack River watershed
has four superfund sites, two power plants, three sewage treatment plants,
and roughly 1,000 ha of landfills within 4 km of the river. The river itself is
brackish with the salinity higher downstream. Tide gates prevent the free mix-
ing of saline waters from the main stem of the river. Highest salinity (12-15 ppt
is found in the main channel closest to the bay, and lowest salinity (<1 ppt) is
found behind tide gates and up tidal creeks where freshwater from combined
sewer overflows and sewage treatment plants are highest (Shin et al., 2013).
The low marshes along the Hackensack are dominated with a tall vigorous
form of S. alterniflora, mixed with bare mudflats; black grass marshes (Jun-
cus gerardii) with stunted Spartina alterniflora. Spartina patens and Distichlis
spicata dominate the high salt marshes. Extensive reed beds dominated by
Phragmites australis are found throughout the Hackensack’s brackish marshes
(Artigas and Pechmann, 2010).

Eleven wetland restoration sites in the Hackensack Meadowlands District
are identified in Figure 18.7a, probably covering less than 300 ha. At least

616 Chapter 18 Wetland Creation and Restoration

23 wetland mitigation sites are located in the district (R. M. Feltes, personal
communication 2014), with the first constructed in 1983. Almost all of the
mitigation conducted in the Meadowlands has been done in kind for impacts
to tidal brackish waters. In-lieu fee agreements were used by the New Jersey
Meadowlands Commission to acquire and enhance ecological functions at Har-
rier Meadow (constructed in 1998), Skeetkill Marsh (1998), Mill Creek Marsh
(1999) and the Secaucus High School Wetland Enhancement Site (SHSWES;
2007). Although Mill Creek Marsh and SHSWES satisfied mitigation obliga-
tions for several different project permits, both public and private, these were
not formal mitigation banks. Mitigation banks in the Meadowlands include
Marsh Resources Meadowlands Mitigation Bank (MRI), Phases 1 and 2. Marsh
Resources planting was completed in 2001 and continues to be monitored
and managed for invasive plants. The 87-ha Richard P Kane Natural Area Mit-
igation Bank, constructed in 2012, was designed largely as tidal Spartina
alterniflora low marsh but has suffered from numerous design and financial
challenges (R. M. Feltes, personal communication, 2014). Also a 20-ha Ever-
green MRI Phase 3 Mitigation Bank was constructed and planted in 2012. The
commission had an agreement with Rutgers University to conduct much more
extensive monitoring of the Harrier Meadow, Skeetkill, and Mill Creek Marsh
mitigation sites.

Mangrove Restoration

Restoring mangrove swamps in tropical regions of the world has some similar charac-
teristics to restoring salt marshes in that the establishment of vegetation in its proper
intertidal zone is the key to success. But that is generally where the similarities end.
Mangrove restoration is more cosmopolitan in that it has been attempted through-
out the tropical and subtropical world (Lewis, 2005); salt marsh restoration has been
attempted primarily on the eastern North American and Chinese coastlines and to
some extent in Europe and the West Coast of North America. Salt marsh restoration
often can rely on waterborne seeds distributing through an intertidal zone; mangrove
restoration often involves the physical planting of trees, although recent work has
shown that these planting often fail (Samson and Rollon, 2008; Lewis, 2009). In
countries such as Vietnam, mangrove declines have been attributed to: the spraying
of herbicides during the Vietnam war; immigration of people to the coastal regions,
leading to cutting of lumber for timber, fuel, wood, and charcoal; and extensive con-
version of mangrove forests to shrimp aquaculture ponds.

Mangroves have being cleared for decades for construction of aquaculture
ponds at unprecedented rates in Vietnam and many other tropical coastlines of the
world (Benthem et al., 1999; Lewis and Brown, 2014). Most of the edible shrimp
sold in the United States and Japan are produced in artificial ponds constructed in

Coastal Wetland Restoration 617

mangrove wetlands in Thailand, Indonesia, and Vietnam. Sold in the United States
and Japan at very low prices, these products are the result of massive destruction of
mangrove forests. More than 100,000 ha of abandoned ponds located in former man-
grove swamps currendy exist in these countries (R. Lewis, personal communication).
In Vietnam, mangroves are being restored and protected to provide coastal protection
and coastal fisheries support. In the Philippines, despite a presidential proclamation
prohibiting the cutting of mangroves, it is estimated that the country still was losing
3,000 ha/yr in the late 1990s (2.4 percent/yr; deLeon and White, 1999). But the
insatiable appetites in the United States, Japan, and several other developed countries
for shrimp continue to cause mangroves to be destroyed. The shrimp ponds last
only about five to six years before they develop toxic levels of sulfur; then they are
abandoned and more mangroves are destroyed. These abandoned ponds present a
challenge for mangrove restoration. Recent efforts to undertake restoration using
Community Based Ecological Mangrove Rehabilitation in Indonesia have proved
successful, and more large-scale restoration of these habitats is under way (Brown
et al. 2014).

Lewis (2005), Lewis and Gilmore (2007), and Lewis and Brown (2014) argue
that common ecological engineering approaches would work best in restoring
mangrove swamps and that more of an analytic approach and less of a “gardening”
approach should be taken. They recommend seven principles to correctly restore
mangroves:

1 . Get the hydrology right.

2. Do not initially build a nursery but carefully determine the reason for a lack
of volunteer mangroves at a proposed restoration site.

3. See if the conditions that prevent natural colonization can be corrected. If
they cannot, pick another site.

4. Examine normal hydrology and topography in reference mangrove swamps
as your model for restored sites.

5. Remember that mangrove swamps do not have flat floors but have subtle
topographic patterns.

6. Gonstruct tidal creeks to facilitate flooding and drainage of tide waters and to
improve fish and invertebrate access to mangroves for harvest by local fishers.

7. Evaluate the costs of restoration early in the project design to make the
project as cost effective as possible.

Two manuals designed to assist with design and construction of successful mangrove
restoration sites (Primavera et al., 2012; Lewis and Brown, 2014) describe what has
been termed “ecological mangrove restoration” (EMR). EMR utilizes a step-wise
approach to engineering large-scale mangrove restoration projects that generally avoid
the wasted time and money of cultivating mangroves in nurseries but instead depend
on natural recruitment of mangrove seeds and propagules to properly restored sites.
Gase Study 6 provides a description of the ambitions and failures of one of the largest
mangrove restorations ever attempted in the world.

618 Chapter 18 Wetland Creation and Restoration

CASE STUDY 6: Restoring Mangroves after the 2004 Indian Ocean
Tsunami

An estimated 230,000 people were killed in late December 2004 as a result of
a massive tsunami around the Indian Ocean caused by an earthquake off the
coast of Sumatra, Indonesia. (See “Mangrove Swamps and the Indian Ocean
Tsunami of December 2004” in Chapter 16: “Wetland Ecosystem Services.”)
The 2004 Indian Ocean tsunami initially caused great interest in restoring man-
grove and other coastal ecosystems to replace areas stricken by the tsunami
as well as to provide coastal protection in the event of future tsunamis or other
tidal surges. Immediately after the cleanup was under way, the whole world was
made aware that the previous destruction of mangrove forests bears some of
the blame for the high loss of human life and cultural impact. Soon after-
ward, mangrove and other coastal vegetation restoration was determined to
be the best approach for local governments to ensure that a similar disaster
would never happen again. Many countries and regions adopted this strategy,
and efforts were undertaken all around the Indian Ocean. The governments of
Malaysia, India, and Indonesia alone promised a total of U.S.$55 million to
replant mangrove forests along their respective coastlands.

Mangrove restoration is really ecological engineering and many of the
restored mangrove forests failed because the plantings occurred in regions
where the tidal and hydrologic conditions are not appropriate (Lewis, 2005,
2010). Follow-up investigations of the success of the extensive plantings of
mangroves after the tsunami have shown little real success, despite the plant-
ing of approximately 30 million mangrove seedlings (LINE!? 2007, 2008). At
least half of all of the plantings failed within just a few years. In fact, Lewis
(2010) declared that there is little evidence that mangroves have ever been
restored on a large scale. He attributed this to two misguided assumptions
regarding mangrove restoration: “(1) mangroves can only be restored by plant-
ing, and (2) sub-tidal mud flats are suitable for planting mangroves, when in
fact they likely never supported a mangrove forest in the first place."

Millions of U.S. dollars were wasted worldwide partially because of the fail-
ure of land managers to recognize the overwhelming importance of self-design
in mangrove recovery and to understand the basic ecology of mangrove veg-
etation. The other negative side of this project is that as we have passed
the 10-year anniversary of the tsunami, interest in restoring mangroves has
almost disappeared. Check (2005) reported that despite many local man-
grove replantings and massive public assistance provided by international
organizations, such as the United Nations, many tropical coastline regions
returned to their old way of destroying mangrove forests for short-time prof-
itable shrimp farms, making the regions even more susceptible than before
to tropical storms and tsunamis.

Coastal Wetland Restoration 619

Delta Restoration

As large rivers connect to the sea, multi-tributary deltas tend to develop, allowing
the river to discharge to the sea in many channels. Many of these rich-soil deltas are
among the most important ecological and economic regions of the world, from the
ancient Nile delta in Egypt to the modern-day Mississippi River Delta in Louisiana.
There should be two major ecological resource goals of delta areas: (1) protecting
and restoring the functioning of the deltaic ecosystems in the context of a geologi-
cally dynamic framework; and (2) controlling pollution from entering the downstream
lakes, oceans, gulfs, and bays. Delta restoration should have this dual emphasis where
possible — ecosystem enhancement of the delta itself and improvement of coastal water
quality downstream. The best strategy for delta restoration when “land building” is a
necessary prerequisite is to restore the ability of the river to “spread out its sediments”
in deltaic form through as wide an area as possible, particularly during flood events
and by not discouraging (or encouraging and even creating) river distributaries. When
river distributaries are not possible on a large scale because of navigation requirements
or population locations, then restoring and creating riverine wedands and construct-
ing river diversions to divert river water to adjacent lands may be the best alternatives
to maximize nutrient retention and sediment retention. In some cases, this involves
the conversion of agricultural lands back to wedands; in other cases, the dikes that
“protect” wildlife protection ponds or retain rivers in their channels only need to be
carefully breached to allow lateral flow of rivers during flood season. See further dis-
cussion of delta restoration for the dual purpose of wedand enhancement and water
quality improvement in Case Study 4: Diverdng the Mississippi River to the Louisiana
Delta in Chapter 19: “Wedands and Water Quality.”

River Restoration

Rivers, well upstream of their deltas with the sea, are being restored around the world
at a rapid rate (Bernhardt et ah, 2005). In the past 10 years, there has been a paradigm
shift in river restoradon. Previously efforts centered on improving in-stream habitats
by remeandering streams and adding physical structures, such as artificial riffles, in the
stream itself (Hart et ak, 2002). Current efforts involve restoradon of the endre rivers
and its corridor in what really is ecological engineering (Palmer et ak, 2014). This
restoradon invariably involves a mix of creating and restoring a combination of fluvial-
and shallow-water wetland systems, often with much of the emphasis on restoring the
riparian edges to the river. Palmer et ak (2005) proposed five criteria for measuring
success of river restoradon:

1 . The design of an ecological river restoration project should be based on a
specified guiding image of a more dynamic, healthy river that could exist at
the site.

2. The river’s ecological condidon must be measurably improved.

3. The river system must be more self-sustaining and resilient to external
perturbadons so that only minimal follow-up maintenance is needed.

620 Chapter 18 Wetland Creation and Restoration

4. During the construction phase, no lasting harm should be inflicted on the
ecosystem.

5. Both pre- and post- assessment must be completed, and data must be made
publicly available.

Now the emphasis is on restoring the entire river ecosystem that involves reconnecting
the river to its floodplain and floodplain wetlands and restoration of the watershed itself
(Mitsch et ah, 2008; Kristensen et ah, 2014). Case Study 7 describes one of the largest
and most complete river restorations in Europe, where restoration involved both the
stream itself and the floodplain wetlands. It is also a site where long-term monitoring
has occurred and allowed significant opportunities for estimating if restoration success
is possible in a decadal time frame.

CASE STUDY 7: Restoring the River Skjern in Denmark

The Skjern River in west-central Jutland drains a watershed of 2,490 km^ and
is the largest in Denmark. In Denmark’s largest drainage project ever, 4,000
ha of wet meadow (Fig. 18.8a) was converted into arable land, and the lower
Skjern was straightened to a fraction of its former meandering self. By the
late 1980s, the river was essentially a straight line to the Ringkobing Fjord
on the North Sea, eliminating thousands of hectares of marshland, mead-
ows, and river habitat (Fig. 18.8b). The channelized river was diked, canals
were built, and pumps were installed to hasten the downstream movement
of water from the land. This public works project cost DDK 30 million (about
U.S.$3.6 million) and was considered a success by the agricultural community
at first as grains could now be grown in the formerly wet region. But the envi-
ronment was paying a heavy price with this artificial river. The self-cleansing
ability of the river was lessened, the downstream fjord was becoming pol-
luted with nutrients and sediments, and the land that was draining began
to subside due to peat oxidation and loss of water — up to 1 m or more in
some locations. The human interference on this river has been described as
“some of the most severe in northern Europe” by the Danish Ministry of Envi-
ronment and Energy (DMEE, 1999). A few short years after the drainage, it
appeared that another drainage project might be necessary, and public fund-
ing was requested. Instead, the Danish parliament (Folketing) passed a Public
Works Act in 1998 by a huge majority that called for the restoration of the lower
Skjern River and earmarked about US$40 million (DDK 254 million) for this
project. The project was implemented in three phases for three reaches of the
river. The river restoration in this case called for the following:

■ Put back the meanders of the river wherever possible.

■ Remove dikes along the river to allow adjacent meadows to be flooded
once again.

Coastal Wetland Restoration 621

■ Define the project area by dikes far away from the river to prevent
flooding of farmland outside of the project area.

Figure 18.8 Restoration of Skjern River and its fioodpiain wetlands in western Den-
mark: (a) meadows and marshlands prior to channelization (1871); (b) channelized
stream (1987) prior to restoration; and (c) stream, meadows, and marshlands almost a
decade after restoration (2011); (d) photo of restored Skjern River and floodplain June
2012, 10 years after restoration, ((a), (b), (c) from DMEE, 1999, and Kristensen et al.,
2014; (d) photo by W. J. Mitsch)

622 Chapter 18 Wetland Creation and Restoration

Figure 18.8 {Continued)

The restoration involved transforming 19 km of channelized river to 26 km
of meandering river (Pedersen et al., 2007a, b). About 2,200 ha of the river
valley wetlands were restored between 1998 and 2002 by removing dikes
and levees that were adjacent to and moving them far away from the river to
protect farmland outside of the project area. Early on, the project was suc-
cessful in substantially increasing the biodiversity of aquatic macrophytes,
invertebrates, amphibians, and mammals such as otters (Pedersen et al.,
2007a). Kristensen et al. (2014) compared in-stream habitats and sedimen-
tation/erosion as well as reconnections with the floodplain for three periods:
(1) year 2000 — before restoration; (2) year 2003 — immediately after restora-
tion; and (3) 2011 — a decade after restoration (Fig. 18.8c, d). They found
no long-term significant changes in in-stream habitats beyond what happened
immediately after restoration in 2003. There was a net erosion on both banks
of the restored stream but net sedimentation in the riverbed; overall, the
restoration resulted in “a slightly wider and shallower river in 2011 than in
2001” (Kristensen et al., 2014). Processes that reshape the channels are
slow and may take centuries. Reconnection between the river and its flood-
plain and the river channel was immediate, and 611 ha of riparian ecosystems
were flooded 10 percent of the time, mostly during the winter; flooding of the
riparian ecosystems was rare (<1 percent) during the growing season of May
through August (Fig. 18. 8d). It may take centuries for lost habitats on the
floodplains, such as islands, backwaters, and oxbow lakes, to develop unless
restoration engineering is employed to restore the lost habitats more quickly
(Kristensen et al., 2014).

Wetland Creation and Restoration Techniques 623

Wetland Creation and Restoration Techniques

Defining Goais

The design of an appropriate wetland or series of wetlands, whether for habitat recre-
ation, the control of non-point source pollution, or wastewater treatment, should start
with forming the overall objectives of the wetland. One view is that wetlands should
be designed to maximize ecosystem longevity and efficiency and minimize cost. The
goal, or a series of goals, should be determined before a specific site is chosen or a
wetland is designed. If several goals are identified, one must be chosen as primary.

Placing Wetlands in the Landscape

In some cases, particularly when sites are being chosen for habitat replacement, many
site choices are available in the landscape to locate a restored or created wetland.
The natural design for a riparian wetland fed primarily by a flooding stream or river
(Fig. 1 8. 9a) allows for flood events of a river to deposit sediments and chemicals on
a seasonal basis in the wedand and for excess water to drain back to the stream or
river. Because there are natural and also often constructed levees along major sections
of streams, it is often possible to create such a wetland with minimal construction.
The wetland could be designed to capture flooding water and sediments and slowly
release the water back to the river after the flood passes, or to receive flooding water
and retain it through the use of flap gates.

Wetlands can be designed as in-stream systems by adding control structures to
the streams or by impounding a distributary of the stream (Fig. 1 8. 9b). Blocking
an entire stream is a reasonable alternative only in headwater streams, and it is not
generally cost-effective or ecologically advisable. This design is particularly vulnerable
during flooding, and its stability might be unpredictable, but it has the advantage
of potentially treating a significant portion of the water that passes that point in the
stream. The maintenance of the control structure and the distributary might mean
making significant management commitments to this design.

A riparian wetland fed by a pump (Fig. 1 8. 9c) creates the most predictable hydro-
logic conditions for the wetland but at an obvious extensive cost for equipment and
maintenance. If it is anticipated that the primary objective of a constructed wetland
is the development of a research program to determine design parameters for future
wetland construction in the basin, then wetlands fed by pumps is a good design. Two
examples of wetlands of this type constructed primarily for research and education are
the Des Plaines River wetlands in northeastern Illinois (Sanville and Mitsch, 1994)
and the Olentangy River wetlands in central Ohio (Mitsch et al., 1998, 2012, 2014).
If other objectives are more important, then the use of large pumps is usually not
appropriate unless the wetland is constrained in an urban setting with no recourse.

Locating several small wetlands on small streams or intercepting ditches in the
upper reaches of a watershed (but not in the streams themselves), rather than creating
fewer larger wetlands in the lower reaches, should be considered (Fig. 18. 9d). The
usefulness of wetlands in decreasing flooding increases with the distance the wetland
is downstream.

Figure I8.9e shows a design involving the creation of a wetland along a stream
to intercept tile drains from agricultural fields. The stream is not diverted, but the

624 Chapter 18 Wetland Creation and Restoration

Figure 18.9 Landscape locations of created and restored wetlands in a riverine setting:

(a) riparian wetland that both intercepts groundwater from uplands but also receives annual
flood pulse from adjacent river; (b) riparian wetland with natural flooding; (c) riparian wetland
with pump; (d) multiple upstream wetlands versus single downstream wetland; and (e) lateral
wetland intercepting groundwater carried by tile drains.

wetlands receive their water, sediments, and nutrients from small tributaries, swales,
and especially tile drains that otherwise would empty straight to the stream. If tile
drains can be located and broken or blocked upstream to prevent their discharge into
tributaries, they can be rerouted to make effective conduits to supply adequate water
to constructed wedands. Because tile drains are often the sources of the highest con-
centrations of chemicals, such as nitrates from agricultural fields, the lateral wetlands
would be an efficient means of controlling certain types of non-point source pollution
while creating a needed habitat in an agricultural setting.

Wetland Creation and Restoration Techniques 625

Site Selection

Several important factors ultimately determine site selection. When the objective is
defined, the appropriate site should allow for the maximum probability that the objec-
tive can be met, that construction can be done at a reasonable cost, that the system
will perform in a generally predictable way, and that the long-term maintenance costs
of the system are not excessive. These 12 factors are elaborated next.

1 . Wetland restoration is generally more feasible than wetland creation.

2. Take into account the surrounding land use and the future plans for the
land.

3. Undertake a detailed hydrologic study of the site, including a determination
of the potential interaction of groundwater with the proposed wetland.

4. Find a site where natural inundation is frequent.

5. Inspect and characterize the soils in some detail to determine their
permeability, texture, and stratigraphy.

6. Determine the chemistry of the soils, groundwater, surface flows, flooding
streams and rivers, and tides that may influence the site water quality.

7. Evaluate on-site and nearby seed banks to ascertain their viability and
response to hydrologic conditions.

8. Ascertain the availability of necessary fill material, seed, and plant stocks and
access to infrastructure (e.g., roads, electricity).

9. Determine the ownership of the land and hence the price.

10. For wildlife and fisheries enhancement, determine if the wetland site is
along ecological corridors, such as migratory fiyways or spawning runs.

11. Assess site access.

12. Ensure that an adequate amount of land is available to meet the objectives.

Creating and Maintaining the Proper Hydrology

The key to restoring and creating wetlands is to develop appropriate hydrologic con-
ditions. Groundwater inflow is often desired because it offers a more predictable and
less seasonal water source. Surface flooding by rivers gives wetlands a seasonal pattern
of flooding, but such wedands can be dry for extended periods in flood-absent peri-
ods. Depending on surface runoff and flow from low-ordered streams can be the
least predictable. Often wetlands developed in these conditions are isolated pools
and potential mosquito havens for a good part of the growing season; their design
should be carefully considered. It is generally considered to be optimum to build
wetlands where they used to be and where hydrology is still in place for the wetland
to survive. But tile drainage, ditches, and river downcutting have often changed local
hydrology from prior conditions. Most biologists have difficulty estimating hydrologic
conditions, while engineers often overengineer control structures that need substantial
maintenance and are not sustainable.

Wetland basins are constructed either by establishing levees around a basin that
may be partially excavated in the landscape or by excavating a depression without

626 Chapter 18 Wetland Creation and Restoration

discharge pipe
/

removabie 2” x 6' stoplogs
for regulating water levels

removable 2” x 6' stoplogs

Figure 18.10 Designs fer control systems for created and restored wetlands including
(a) drop pipe, (b) flashboard riser, and (c) fuii-round riser. (From Massey, 2000)

constructing any levee. Construction engineers often note that if they use excavated
soil for a levee, they can save large sums of money because excavation is often the
largest cost of wetland restoration or creation. This levee construction is usually not
a good idea, because levees are bound to have problems with leakage and, in many
parts of the world, burrowing animals like muskrats {Ondatra zibethicus).

Some sort of control structure is often needed at the outflow of the wetland basin,
whether there is a levee or not. The control devices are the outflow of the wetland.
Three such control devices are shown in Figure 18.10: (1) drop pipes, (2) flashboard
risers, and (3) full-round riser (combination of drop pipe and flashboard riser).

Each device has its own advantages and disadvantages. Drop pipes are the least
flexible because they do not allow water-level manipulation. A flashboard riser is more
flexible but can be easily vandalized. Full-round risers are a little more secure and can
be designed for control of beavers, but they are a little more expensive. In the last
two cases, the outflow risers include removable stoplogs that allow manual changes in
water level. This option is desirable when the exact hydrology of the wetland basin is
not known, because it allows flexibility.

But these types of control devices have several disadvantages. They require
occasional maintenance, if only for removing accumulation of plant debris and
resetting stoplogs. Also, stoplog removal is a favorite pastime of vandals. Control
devices such as risers are also favorite locations for nature’s ecological engineer — the

Wetland Creation and Restoration Techniques 627

beaver Castor canadensis — to provide its idea of water management, usually creating
blockages that can raise the water level by a meter or more, changing the vegetation
patterns dramatically.

The best design situation is when the local topography allows the wetland to be
naturally flooded without control devices, but this opportunity is rarely available. For
a reliable source of water, groundwater is generally less sensitive to seasonal highs and
lows than is surface water. Also, a wetland fed by groundwater invariably has better
water quality and generally fewer sediments that will eventually fill the wedands.

Soils

Choice of the site for wetland creation and restoration is often limited by property
ownership. If a choice exists, a wetland that is restored on former wetland (hydric)
soils is much preferred over one constructed on upland soils. Hydric soils develop
certain color and chemical patterns, because they have spent long periods flooded
and thus under anaerobic conditions. The soil color is mostly black in mineral hydric
soils, because iron and manganese minerals have been converted to reduced soluble
forms and have leached out of the soil (see Chapter 5 : “Wetland Soils” ) . In most cases,
developing wetlands on hydric soils has three three advantages:

1 . Hydric soils indicate that the site may still have or can be restored to
appropriate hydrology.

2. Hydric soils may be a seed bank of wetland plants still established in the soil.

3 . Hydric soils may have the appropriate soil chemistry for enhancing certain
wetland processes. For example, mineral hydric soils generally have higher
soil carbon than do mineral nonhydric soils. This soil carbon, in turn,
stimulates wetland processes such as denitrification and methane production.

Otherwise, it is possible to create wetlands on upland soils. In the long run, those
soils will develop characteristics typical of hydric soils, such as higher carbon content
and seed banks (see Case Study 8).

CASE STUDY 8: Hydric Soil Development and Chemical Sequestration
in Created Wetlands

It was not well known until recently how long It would take upland soils to
develop wetland conditions; it was thought to be over a decade or even a cen-
tury, depending on the soil types and the hydrology. In the study of the two
wetlands at the Olentangy River Wetland Research Park in Ohio that were cre-
ated in 1994, hydric soil conditions were shown to develop after only two
years of continued flooding (Mitsch et al., 2005). Before the basins were
first flooded, the most prevalent hue was lOYR, and the value/chroma soil
color varied between 3/3 and 3/4. The chroma of 3 to 4 indicates nonhydric
soils (see description of these soil color terms in Chapter 5). In 1995, about
18 months after flooding, chromas of 3 or less were common (median = 3/2).

Figure 18.11 Soil organic matter development in two 1-ha created experimental wetlands over a 12-year period at the Olentangy
River Wetland Research Park. Left: Spatial distribution maps of soii organic matter for Wetiands 1 and 2 shown for: (a) 1993,
before water was added but after basins were excavated; (b) 1995, 16 months after pumping began; and (c) 2004, 10 years after
fiooding began. Right: Fh'equency distribution curves for soii organic matter in two experimentai wetiands in 1993, 1995, and 2004
for (d) pianted Wetiand 1 and (e) unplanted Wetland 2. Wetlands were excavated in nonhydric alluvial floodplain soils in 1993, and
water has been continuously pumped into both wetlands since March 4, 1994. (Fk'om Anderson et al., 2005)

628

Wetland Creation and Restoration Techniques 629

The mean value in the surface samples was 3/2; subsurface sample median
values were 4/2. Chromas started to consistently be of 2 or below in 1996,
two years after flooding began. As of 2006, 13 years after the soils were
flooded, almost all samples in these experimental wetlands show chromas of
2 or less In the surface sediments.

Lack of organic carbon in the soil is often described as a shortcoming
of wetland creation projects. The organic content of the upper soils in these
experimental wetlands increased steadily over the first decade, from 1994
when water was first added (Fig. 18.11; Table 18.3). The organic content of
the surface (0-8 cm depth) soils increased from 5.3 ± 0.1 percent in 1993
(before water was added), to 6.1 ± 0.1 percent in 1995 (18 months after cre-
ation), to 8.9 ± 0.2 percent in 2004 (10 years after the wetland was created)
(Anderson et al., 2005). Total carbon increased from 1.57 ± 0.04 percent in
1993, to 2.06 ± 0.12 percent in 1995, and to 3.76 ± 0.12 percent in 2004.
In other words, the organic content and total carbon of these wetland sur-
face sediments increased by 67 percent and 139 percent, respectively, over a
decade. While most of the carbon increase is believed to be due to algal and
macrophyte productivity, it is clear from other studies at these wetlands (Wu
and MItsch, 1998; Liptak, 2000; Tuttle et al., 2008) that a substantial amount
of the carbon accumulation is probably due to inorganic calcium carbonate/
calcite (CaCOg) (See inorganic carbon accumulation in Table 18.3) that pre-
cipitates at high rates in the growing season due to water column productivity
in these wetlands. A rule of thumb from this study on newly created wetlands:
Organic content in surface soils in newly created wetlands increases by about
1 percentage point every three years.

Table 18.3 Mean carbon and nutrient accumulation rates in

two experimental wetlands at the Olentangy River Wetland

Research Park, Ohio, 1994 to 2004

Parameter Mean Annual Accumulation Rates g m"^ yr~^

Total carbon

181-193

Organic carbon

152-166

Inorganic carbon

23-26

Total nitrogen

16.2-16.6

Total phosphorus

3.3-3.5

Total calcium

80.8-86.3

Source: Anderson and MItsch (2006).

It has been argued that upland soils often do not allow the development of a
major diversity of plant communities but often become Typha marshes instead because
of the absence of seed banks. This domination by Typha is as much due to the fact

630 Chapter 18 Wetland Creation and Restoration

that uplands converted to wedands have often been used for agriculture for many
years and are thus quite eutrophic. The high-nutrient conditions invariably lead to
high -productivity, low-diversity systems. Again, the main advantage of using hydric
soils in wetland restoration and creation is that they are indicators of appropriate
hydrologic conditions.

Introducing Vegetation

The species of vegetation types to be introduced to created and restored wetlands
depend on the type of wetland desired, the region, and the climate as well as the
design characteristics described previously. Table 18.4 summarizes some of the plant
species used for wetland creation and restoration projects, mostly in the United States.
Vymazal (2013) found in a literature search of 643 surface-flow constructed wetlands
described in 43 countries that 150 species have been used. The most commonly used
genera are Typha, Scirpus, Schoenoplectus, Phraprmites, Juncus, and Eleocharis. More
details are given in Chapter 19 on plants used for water quality improvement.

Freshwater Marshes

Common plants used for freshwater marshes include bulrush {Scirpus spip. and Schoeno-
plectus spp.), cattails {Typha spp.), sedges {Carex spp.), and floating-leaved aquatic
plants such as white water lilies {Nymphaea spp.) and spatterdock {Nuphar spp.).
Submerged plants are not common in wetland design, and their propagation is often
hampered by turbidity and algal growth in the early years of wetland development.

Coastal Marshes

For coastal salt marshes, Spartina alterniflora is the primary choice for coastal marsh
restoration in the Eastern United States. Both Spartina townsendii and S. an^lica have
been used to restore salt marshes in Europe and in China. The details of successful
coastal wetland creation are site specific, but several generalizations seem to be valid
in most situations.

Forested Wetlands

Forested wetland restoration and creation usually involve the establishment of
seedlings. In the southeastern United States, deciduous hardwood species typical
of bottomland forests are planted. They include nuttall oak {Quercus nuttallii),
cherrybark oak {Q. falcata var. pagodifoUa)^ willow oak {Q. phellos)^ water oak
{Q nip'ra), cottonwood {Populus deltoides), sycamore {Platanus occidentalis), green
ash {Fraxinus pennsylvanica), sweetgum {Liquidambar styracifula), and pecan {Cary a
illinoensis) . There is less use of deepwater plants such as bald cypress {Taxodium
distichum) and water tupelo {Nyssa aquatica), although Taxodium spp. was once
the dominant genus of introduced species in many wetland restorations in Florida
(Clewell, 1999). In Florida, a wide variety of wetland oaks, bays, gums, ashes, and
pines are also used in forested wetland restoration.

Table 18.4 Selected plant species planted in created and restored wetlands

Scientific Name

Common Name

Scientific Name

Common Name

Freshwater Marsh — Emergent

Acorus calamus

sweet flag

Pontederia cordata

pickerelweed

Cladium jamaicense

sawgrass

Sagittaria rigida

duck potato

Carex spp.

sedges

Sagittaria latifoiia

duck potato; arrowhead

Eleocharis spp.

spike rush

Saururus cernuus

lizard's tail

G/ycer/a spp.

manna grass

Schoenoplectus

soft-stem bulrush

Hibiscus spp.

rose maiiow

tabernaemontani*

Iris pseudacorus

yeiiow iris

Scirpus acutus

hard-stem bulrush

Iris versicolor

biue iris

Scirpus americanus

three-square bulrush

Juncus effusus

soft rush

Scirpus cyperinus*

Woolgrass

Leersia oryzoides

rice cutgrass

Scirpus fluviatilis

river bulrush

Panicum virgatum

switchgrass

Sparganium eurycarpum

giant bur reed

Peltandra virginica

arrow arum

Spartina pectinata

prairie cordgrass

Phalaris arundinacea

reed canary grass

Typha angustifoUa*

narrow-leaved cattail

Phragmites australis*

giant reed

Typha latifoiia*

wide-leaved cattail

Polygonum spp.

smartweed

Zizania aquatic

wild rice

Freshwater Marsh — Submerged

Ceratophyllum demersum

coontail

Potamogeton pectinatus

Sago pondweed

Elodea nuttallii

waterweed

Vallisneria spp.

wild celery; tape grass

Myriophyllum aquaticum

milfoil

Najas guadaiupensis*

Southern naiad

Freshwater Marsh — Floating

Azolla caroliniana

water fern

Nuphar luteum

spatterdock

Eichhornia crassipes*

water hyacinth

Pistia stratiotes

water lettuce

Hydrocotyle umbellata

water pennywort

Salvinia rotundifolia

floating moss

Lemna spp.

duckweed

Wolffia sp.

water meal

Nymphaea odorata

fragrant white water iiiy

Bottomlands/Forested Wetland

Acer rubrum

red mapie

Gordonia lasianthus

loblolly bay

Acer floridanum

Fiorida mapie

Liquidambar styracifula

sweetgum

Acer saccharinum

silver maple

Platanus occidentalis

sycamore

Alnus spp.

alder

Populus deltoids

cottonwood

Carya illinoensis

pecan

Quercus falcata var. pagodifolia

cherrybark oak

Celtis occidentalis

hackberry

Ulmus Americana

American elm

Cephalanthus occidentalis

buttonbush

Quercus nigra

water oak

Comas stolonifera

red-osier dogwood

Quercus nuttaliii

Nuttall oak

Fraxinus caroliniana

water ash

Quercus pheilos

willow oak

Fraxinus pennsylvanica

green ash

Salix spp.

willow

Deepwater Swamp

Nyssa aquatica

swamp tupelo

Taxodium distichum

bald cypress

Nyssa sylvatica var. bifiora

black gum

Taxodium distichum var. nutans

pond cypress

aka Taxodium distichum var.

imbricarium

631

632 Chapter 18 Wetland Creation and Restoration

Table 18.4 (Continued)

Scientific Name

Common Name

Scientific Name

Common Name

Salt Marsh

Distichlis spicata

spike grass

Spartina foliosa

cordgrass (Western U.S.)

Salicornia sp.

Saltwort

Spartina patens

salt meadow grass

Spartina alterniflora
Spartina anglica

cordgrass (Eastern U.S.)
cordgrass (Europe; China)

Spartina townsendii

cordgrass (Europe)

Mangrove Swamp

Rhizophora mangle
Avicennia germinans

red mangrove
black mangrove

Laguncularia racemosa

white mangrove

•Commonly planted in treatment wetlands.

Wetland Planting Techniques

Plants can be introduced to a wetland by transplanting roots, rhizomes, tubers,
seedlings, or mature plants; by broadcasting seeds obtained commercially or from
other sites; by importing substrate and its seed bank from nearby wetlands; or by
relying completely on the seed bank of the original and surrounding site. If planting
stocks rather than site seed banks are used, it is most desirable to choose plants from
wild stock rather than nurseries because the former are generally better adapted to the
environmental conditions they will face in constructed wetlands. The plants should
come from nearby if possible and should be planted within 36 hours of collection.
If nursery plants are used, they should be from the same general climatic conditions
and should be shipped by express service to minimize losses. Marshes should be
planted at densities to ensure rapid colonization, adequate seed source, and effective
competition with undesirable plants, such as Typha spp. Specifrcally, this could mean
introducing from 2,000 to 5,000 plants/ha.

For emergent plants, the use of planting materials with stems of at least 20 to
30 cm is recommended, and whole plants, rhizomes, or tubers rather than seeds have
been most successful. In temperate climates, both fall and spring planting times are
possible for certain species, but spring plantings are generally more successful, because
it is a better time to minimize destructive winter grazing of plants by migratory animals
and the uprooting of the new plants by ice.

Transplanting plugs or cores (8-10 cm in diameter) from existing wedands is
another technique that has been used with success, for it brings seeds, shoots, and
roots of a variety of wetland plants to the newly restored or created wetland.

If seeds and seed banks are used for wetland vegetation, several precautions must
be taken. The seed bank should be evaluated for seed viability and species present. The
use of seed banks from other nearby sites can be an effective way to develop wetland
plants in a constructed wetland if the hydrologic conditions in the new wetland are
similar. Seed bank transplants have been successful for many different species, includ-
ing sedges {Carex spy) .) ^ Sapiittaria sp., Scirpus acutus, S. validus, and Typha spp. The
disruption of the wetland site where the seed bank is obtained must also be considered.

When seeds are used directly to vegetate a wetland, they must be collected when
they are ripe and stratified if necessary. If commercial stocks are used, the purity of

Wetland Creation and Restoration Techniques 633

the seed stock should be determined. The seeds can be added with commercial drills
or by broadcasting from the ground, watercraft, or aircraft. Seed broadcasting is most
effective when there is little to no standing water in the wetland.

Natural Succession versus Horticulture

To develop a wetland that ultimately will be a low-maintenance one, natural suc-
cessional processes need to be allowed to proceed. The best strategy is usually to
introduce, by seeding and planting, as many native choices as possible to allow natural
processes to sort out the species and communities in a timely fashion. Wetlands cre-
ated or restored by this approach are called self-design wetlands. Providing some help to
this selection process (e.g., selective weeding) may be necessary in the beginning, but
ultimately the system needs to survive with its own successional patterns unless signif-
icant labor-intensive management is possible. A somewhat different approach, called
designer wetlands^ occurs when specifted plant species are introduced, and the success
or failure of those plants is used as indicators of success or failure of that wetland. This
is akin to horticulture.

An important general consideration of wetland design is whether plant material
is going to be allowed to develop naturally from some initial seeding and planting or
whether continuous horticultural selection for desired plants will be imposed. W. E.
Odum (1987) suggested, “In many freshwater wetland sites it may be an expensive
waste of time to plant species which are of high value to wildlife. It may be wiser to
simply accept the establishment of disturbance species as a cheaper although somewhat
less attractive solution.” As described above, Samson and Rollon (2008) and Lewis
(2009) found that planting of mangrove seedlings was often a big waste of time and
resources. The successful salt marsh restoration at Delaware Bay described in Case
Study 4 (Teal and Weinstein, 2002; Hinkle and Mitsch, 2005) did not require any
seeding or planting. Reinartz and Warne (1993) found that the way vegetation is
established can affect the diversity and value of the mitigation wetland system. Their
study showed that early introduction of a diversity of wetland plants may enhance the
long-term diversity of vegetation in created wetlands. The study examined the natural
colonization of plants in 11 created wetlands in southeastern Wisconsin. The wetlands
under study were small, isolated, depressional wetlands. A two-year sampling program
was conducted for the created wetlands, aged one to three years. Colonization was
compared to five seeded wedands where 22 species were introduced. The diversity and
richness of plants in the colonized wetlands increased with age, size, and proximity to
the nearest wetland source. In the colonized sites, Typha spp. comprised 15 percent of
the vegetation for one-year wetlands and 55 percent for three-year wedands, with the
possibility of monocultures of Typha spp. developing over time in colonized wedands.
The seeded wetlands had a high species diversity and richness after two years. Typha
cover in these sites was lower than in the colonized sites after two years.

Another study where the effects of plandng versus not planting have been
observed for several years was at two 1-ha experimental wedands in central Ohio
(Table 18.5; See also Case Study 8 above). One wetland was planted with 2,500
plants representing 13 macrophyte species; the other was left as a naturally colonizing
control wedand. In essence, both wedands were different degrees of self-design

634 Chapter 18 Wetland Creation and Restoration

Table 18.5 Summary of vegetation species richness in two 1-ha experimental wetlands in
centrai Ohio through 17 years. Both wetiands were created in 1994. The “planted wetland”
(Wl) was planted with 2,500 individuai piants representing 13 native wetiand species. W2 is
unpianted controi

Wetland
Age Year

Number of
Species

Number of
Wetland
Species

Number of
Planted
Species

Number of
Woody
Species

Number of
Invasive
Species

Total

1996

1998

2008

101

116

2010

118

Source: Mitsch etal., 2012 updated.

because there were no expectations as to what the ultimate cover would be and
there was no “gardening” to get to any endpoint. After three years, both wedands
were principally dominated by soft-stem bulrush Schoenoplectus tabernaemontani (=
Scirpus validus) and were thought to be similar (Mitsch et ah, 1998). After six years,
however, several communities of vegetation continued to exist in the planted basin,
but a highly productive monoculture of Typha dominated the unplanted basin where
it did not have any competition from planted vegetation (Mitsch et ah, 2005). By
2013, both wedands had mostly converged on vegetadon cover of Typha spp. The
wedands did have a few differences in wedand funcdon that persisted a decade or
more after plandng that could be traced to effects of the inidal plandng (Mitsch et ah,
2005, 2012, 2014). And 9 of the 13 planted species were still in the planted wedand
20 years after plandng (only 2 “jumped over” to the unplanted basin), although most
were not common (Mitsch et ah, 2014). Both carbon sequestradon and methane
emissions were consistently higher in the naturally colonizing (unplanted) wetland
due to its higher producdvity (Nahlik and Mitsch, 2010; Sha et ah, 2011; Mitsch
et ah, 2012, 2014; Bernal and Mitsch, 2013) whereas other ecological indicators,
such as macrophyte community diversity, were almost always higher in the planted
wedand. Planting at first appeared to have litde to no effect on water quality but
analyses of 15 years of data of nutrient fluxes showed that the plandng enhanced
overall retendon of phosphorus but reduced the retendon of total nitrogen (see also
Case Study 5 in Chapter 19: “Wedands and Water Quality”).

If plant diversity is desired, plandng makes sense. If producdvity and carbon
sequestradon are desired, it may be a waste of effort to plant unless there are no sources
of plant propagules (e.g., seed banks or inflowing rivers). In either case, a long-term
effect on ecosystem function caused by introducing plants appeared to linger 20 years
after a wedand is planted, but that effect is overshadowed by an overwhelming impact
of natural propagule inputs and self-design.

Exotic or Undesirable Plant Species

In some cases, certain plants are viewed as desirable or undesirable because of their
value to wildlife or their aesthedcs. Reed grass (Phrajjmites australis) is often favored

Estimating Success 635

in constructed wetlands in Europe, and there is real concern for reed die-back around
lakes and ponds in Europe. But reed grass is considered an invasive, undesirable plant
in much of eastern North America, particularly in coastal freshwater and brackish
marshes (Philipp and Field, 2005). Spartina alterniflorais the desired endpoint of salt
marsh restoration on the East Coast of North America but is considered an invasive
nuisance on the West Coast and now in China.

Some plants are considered undesirable in wetlands because they are aggressive
competitors. In many parts of the tropics and subtropics, the floating aquatic plants
water hyacinth {Eichhornia crassipes) and alligator weed {Alternanthera- philoxeroides)
are considered undesirable and, in eastern North America, particularly around the
Great Lakes, the emergent purple loosestrife {Lythrum salicaria) is considered an
undesirable alien plant in wetlands. Throughout the United States, cattail ( Typha spp. )
is championed by some and disdained by others, because it is a rapid colonizer but
is of limited wildlife value. In other parts of the world, Typha is considered a per-
fectly acceptable plant in restored wetlands. In New Zealand, several species of willow
{Salix) are invading marshes and other wetlands, and programs to eradicate them
are common.

Estimating Success

There has not been a lot of positive analysis of created and restored wedands in the
literature, despite the fact that many good approaches have been developed. It is our
belief that wetlands can (and should) be created and restored. The problem is more
that those with the right understanding of wetland function are not the ones with the
engineering consulting contracts creating and restoring them.

Few satisfactory methods are available to determine the success of a created or
restored wetland or even a mitigation wetland created to replace the functions lost with
the original wetland. Figure 1 8. 1 illustrated conceptually how it should be done for
replacement wedands. It is clear from several studies of created and restored wedands
that some cases are successes, but there are sdll far too many examples of failures of cre-
ated and restored wetlands to meet expectations. Examples of such studies done in the
1990s for southern Florida, northeastern Illinois, and Ohio are shown in Figure 18.12
and Table 18.6. In some cases, expectadons were unreasonable; in other cases, the
original wedand should not have been lost to begin with. Where expectadons are eco-
logically reasonable, there is optimism that wetlands can be created and restored and
that wedand function can be replaced.

A more recent analysis by Moreno-Mateos et al. (2012) using a meta-analysis
of 621 wedand sites found that biological structure (mosdy esdmated from plant
communities) and biogeochemical funcdon (mosdy esdmated by carbon accumu-
ladon in the wedand soils) were 26 and 23 percent lower, respecdvely, in restored
wedands than in reference wedands. They also found that large wedands (>I00 ha),
tropical/temperate wetlands, and riverine wetlands fared better than did small wet-
lands, cold-climate wetlands, and depressional wetlands, respecdvely.

Hopple and Craft (2013) compared four restored and four natural wetlands in the
glaciated northwest Indiana, including restored and natural wedand sites in the mosdy
drained Kankakee Marsh region (see Chapter 3) and found that, after 10 years, the

636 Chapter 18 Wetland Creation and Restoration

5 _

I Successful
I Limited success
I I Failure
I I Incomplete

■s 50 I-

I Vegetation success

■ Different vegetation
than anticipated

I I Excessive open virater

I I Insufficient water

Figure 18.12 Evaluation of wetland mitigation projects in two regions of the United States:
(a) 40 mitigation projects in south Florida invoiving wetiand creation, mitigation, and preset^
vation. The average age of the projects was less than three years. “Successful” meant that
the project met all of its stated goals, whereas “failure” meant that few goals were met,
or the created/restored wetland did not have functional equivalency to a reference wet-
land; (b) 128 wetland mitigation sites required by 61 permits in the six-county region around
Chicago, liiinois. The permits were issued between 1990 and 1994, and this study began in
1996. ((a) From Erwin, 1991; (b) from Gaiiihugh and Rogner, 1998)

restored wetlands and the natural wetlands had similar plant richness (33.8 + 2.3 ver-
sus 27 + 6.4 species, respectively) and Floristic Quality Assessment Indices. They con-
cluded that “the comparable plot and site diversity of restored wetlands is attributed
to the use of management tools (such as seeding, prescribed burning, and herbicidal
treatments) during restorations that enhance species richness and diversity and shorten
the time required for the plant community of restored wetlands to converge with levels
in natural wetlands.”

Estimating Success 637

Table 18.6 Permit requirements and compliance for five replacement wetlands investigated
in Ohio

Location, County
in Ohio

Wetland Area (ha)

Percentage of Required Area Replaced (%)

Lost

Required

Happened

Location, County
in Ohio

Delaware

3.7

5.4

~4.0

Franklin

15.0

28.0

3.2

Gallia

0.5

0.8

0.7

Jackson

4.8

7.2

7.5

105

Portage

0.4

0.6

100

Total

24.4

42.0

-16.0

Source’. Wilson and Mitsch (1996).

It appears that wetland creation and restoration is being done in a more intelligent
fashion now than it was 20 years ago. The earlier spotty record is due, in our opinion,
to three factors:

1 . Litde understanding of wetland function by those creating and restoring the
wetlands

2. Provision of insufficient time for the wedands to develop

3. A complete lack of recognition or underestimation of the self-design capacity
of nature by biologists and engineers alike

Understanding wetlands enough to be able to create and restore them requires
substantial training in plants, soils, wildlife, hydrology, water quality, and engineer-
ing. Replacement projects and other restorations involving freshwater marshes need
enough time, closer to 15 or 20 years than to 5 years, before success is apparent.
Restoration and creation of forested wetlands, coastal wetlands, or peatlands may
require even more time. Pearland restoration could take decades or more. Forested
wetland restoration generally takes a lifetime. Finally, we should recognize that nature
remains the chief agent of self-design, ecosystem development, and ecosystem main-
tenance; humans are not the only participants in these processes. Sometimes we refer
to these self-design and time requirements for successful ecosystem restoration and
creation as invoking “Mother Nature and Father Time” (Mitsch and Wilson, 1996;
Mitsch et al., 1998, 2012).

Both wetland scientists and wetland engineers need to learn each other’s trade
for the field and the wetlands to be successful (Mitsch, 2014). Wetland science will
continue to make significant contributions to the process of reducing our uncertainty
about predicting wetland success. Wetland creation and restoration need to become
part of an applied ecological science, not a technique without theoretical underpin-
nings. Scientists need to use quantitative and carefully designed experiments to make
the connections between structure, such as vegetation density and diversity, and and
function, such as productivity, wildlife use, organic sediment accretion, and nutrient
retention. Engineers and managers need to recognize that systems that emphasize the
role of self-design and sustainable structures are more ecologically viable in the long
run than are heavily managed systems.

638 Chapter 18 Wetland Creation and Restoration

As Mitsch (2014) summarized:

Ecosystem restoration, as currently practiced throughout the world, is done by
practitioners who have little experience in design (scientists study systems, they do
not design systems) and by engineers who do not appreciate the capabilities of
ecosystems to self-design (engineering is a field devoted to removing uncertainty and
controlling natural processes). The approach of many restorations is restoration by
committee — that results in projects that are less successful than anticipated or are
overdesigned by engineers with unsustainable technology.

Summary Principles

Seven general principles of ecological engineering that apply to the creation and
restoration of wetlands are outlined next (Mitsch and Jorgensen, 2004):

1 . Design the system for minimum maintenance and a general reliance on
self-design.

2. Design a system that utilizes natural energies, such as the potential energy of
streams, as natural subsidies to the system.

3. Design the system with the hydrologic and ecological landscape and climate
in mind.

4. Design the system to fulfill multiple goals, but identify at least one major
objective and several secondary objectives.

5. Give the system time.

6. Design the system for function, not form.

7. Do not overengineer wetland design with rectangular basins, rigid structures
and channels, and regular morphology.

Many other principles can be invoked, but these are good starting points. Zedler
(2000a) had suggested ten ecological principles that should be applied to wetland
restoration, and they fit well with the preceding seven principles:

1 . Landscape context and position are crucial to wetland restoration. See design
principle 3 in the previous list. Wetlands are always a function of the
watershed and ecological setting in which they are placed.

2. Natural habitat types are the appropriate reference systems. This suggests that
while we may know how to build ponds, for example, are those the natural
habitats of the area, even if they do increase waterfowl.^

3. The specific hydrologic regime is crucial to restoring biodiversity and function.
See design principles 2 and 3. In many cases, such as the Florida Everglades,
the restoration is being done in the face of a massive change in the
landscape’s hydrologic character.

4. Ecosystem attributes develop at different paces. Give the system time; see
design principle 5. Hydrology develops quickly, vegetation develops over
several years, and soils develop over decades. Yet we are quick to review and
criticize created and restored wetlands after a couple of years.

References 639

5. Nutrient supply rates affect biodiversity recovery. There are low-nutrient and
high-nutrient wetland systems (see Chapter 7: “Wetland Vegetation and
Succession”). Low-nutrient wedands are often more difficult to create or
restore. With very few exceptions, we live in a eutrophic landscape. (The
Okavango Delta in Botswana might be one.) High-nutrient inflows cause
wetlands to go for power, often instead of diversity.

6. Specific disturbance regimes can increase species richness. This can clearly be
the case if we allow the word disturbance to include flood pulses, fire, and
even tropical storms.

7. Lack of seed banks and dispersal can limit recovery of plant species richness.
This is why restoring wetlands with seed banks can be so important.
Another solution is to have a hydrologically or biologically “open” system
with a multitude of inputs of propagules (plants, animals, microbes) more
likely.

8. Environmental conditions and life-history traits must be considered when
restorinp! biodiversity.

9. Predicting wetland restoration begins with succession theory. Again, design
principle 5 says we need to give the system time. Ecological succession
cannot be accelerated without other consequences. This also supports our
contention that one must understand wetland science first before
attempting to create and restore wetlands.

10. Genotypes influence ecosystem structure and function. This is an important
but often overlooked principle about wetland restoration. Species are not
the same everywhere. This has been shown in common garden experiments
on Spartina alterniflora (Seliskar, 1995) and freshwater rush Juncus effusus
(Weihe and Mitsch, 2000). A brackish/freshwater wetland plant with
several genotypes that has invaded many natural and restored wedands in
the United States is Phra^mites australis.

Recommended Readings

Carey, J. 2013. Architects of the Swamp. Scientific American 309 (6): 74-79.

Mitsch, W. J. 2013. Wedand Creadon and Restoradon. Encyclopedia of Biodiversity .,
2nd ed. pp. 367-383. S. Levin, ed. Amsterdam: Elsevier.

Weinstein, M. P. and J. W. Day (eds.) 2014. Restoration Ecology in a Sustainable
World. Special Issue of Ecolojjical Enjjineerinjj 6S\ 1-158.

References

Alphin, T. D., and M. H. Posey. 2000. Long-term trends in vegetadon dominance
and infaunal community composition in created marshes. Wetlands Ecology and
Management 8: 317-325.

Anderson, C. J., W. J. Mitsch, and R. W. Nairn. 2005. Temporal and spadal devel-
opment of surface soil condidons in two created riverine marshes. Journal of
Environmental Quality 34: 2072-2081.

640 Chapter 18 Wetland Creation and Restoration

Anderson, C. J., and W. J. Mitsch. 2006. Sediment, carbon, and nutrient accumulation
at two 10-year-old created riverine marshes. Wetlands 26: 779-792.

Artigas, F., and I. C. Pechmann. 2010. Balloonimagery verification of remotely sensed
Phra^mites expansion in an urban estuary of New Jersey, USA. Landscape

and Urban Planning 9S-. 105-112.

Atkinson, R. B., J. E. Perry, and J. Cairns, Jr. 2005. Vegetation communities of
20-year-old created depressional wedands. Wetlands Ecology and Management 13:
469-478.

Benthem, W., L. P. van Lavieren, and W. J. M. Verheugt. 1999. Mangrove rehabilita-
tion in the coastal Mekong Delta, Vietnam. In W. Streever, ed.. An International
Perspective on Wetland Rehabilitation. Kluwer Academic Publishers, Dordrecht,
the Netherlands, pp. 29-36.

Bernal, B., and W. J. Mitsch. 2013. Carbon sequestration in two created riverine wet-
lands in the Midwestern United States. Journal of Environmental Quality 42:
1236-1244.

Bernhardt, E. S., M. A. Palmer, J. D. Allen, G. Alexander, K. Barnas, S. Brooks, J. Carr,

S. Clayton, C. Dahm, J. Follstad-Shah, D. Galat, S. Gloss, P. Goodwin, D. Hart,
B. Hassett, R. Jenkinson, S. Katz, G. M. Kondolf, P. S. Lake, R. Lave, J. L. Meyer,

T. K. O’Donnell, L. Pagano, B. Powell, and E. Sudduth. 2005. Synthesizing U.S.
river restoration efforts. Science 308: 636-637.

Blake, N. M. 1980. Land into Water — Water into Land: A History of Water Manage-
ment in Florida, University Press of Florida, Gainesville, 344 pp.

Bousquin, S. G.. ed. 2014. Kissimmee River restoration project. Special section of
Restoration Ecology 22: 345M:34.

Bradshaw, A. D. 1996. Underlying principles of restoration. Canadian Journal of
Fisheries and Aquatic Sciences 53 (Suppl. 1): 3-9.

Broome, S. W., E. D. Seneca, and W. W. Woodhouse, Jr. 1988. Tidal salt marsh
restoration. Aquatic Botany 32: 1-22.

Brown, B., R. Fadilla, Y. Nurdin, 1. Soulsby and R. Ahmad. 2014. Gommunity based
ecological mangrove rehabilitation (GBEMR) in Indonesia. SAPIENS 7(2):
53-64.

Galloway, J. C., and J. B. Zedler. 2004. Restoration of urban salt marshes: Lessons
from southern Galifornia. Urban Ecosystems 7 \ 107-124.

Gheck, E. 2005. Roots of recovery. Nature 438: 910-911.

Ghung, G. H. 1982. Low marshes, China. In R. R. Lewis 111, ed.. Creation and
Restoration of Coastal Plant Communities. CRC Press, Boca Raton, FL, pp.
131-145.

Chung, C. H. 1989. Ecological engineering of coastlines with salt marsh plantations.
In W. J. Mitsch and S. E. Jorgensen, eds. Ecological Engineering: An Introduction
to Ecotechnology . John Wiley 8e Sons, New York, pp. 255-289.

Clewell, A. F. 1999. Restoration of riverine forest at Hall Branch on phosphate-mined
land, Florida. Restoration Ecology 7: 1-14.

Craft, C., S. Broome, and C. Campbell. 2002. Fifteen years of vegetation and
soil development after brackish-water marsh creation. Restoration Ecology 10:
248-258.

References 641

Dahl, T. E. 2006. Status and trends of wedands in the conterminous United States,
1998 to 2004. U.S. Department of the Interior, Fish and Wildlife Service, Wash-
ington, DC. 112 pp.

Danish Ministry of Environment and Energy (DMEE). 1999. The Skjern River
Restoration Project, DMEE and National Forest and Nature Agency. Copen-
hagen, Denmark, 32 pp.

deLeon, R. O. D., and A. T. White. 1999. Mangrove rehabilitation in the Philip-
pines. In W. Streever, ed. , An International Perspective on Wetland Rehabilitation.
Kluwer Academic Publishers, Dordrecht, the Netherlands, pp. 37-42.

Dorcas, M. E., J. D. Willson, R. N. Reed, R. W. Snow, M. R. Rochford, M. A. Miller,
W. E. Meshaka, P. T. Andreadis, F. J. Mazzotti, C. M. Romagosa, and K. M. Hart.
2012. Severe mammal declines coincide with proliferation of invasive Burmese
pythons in Everglades National Park. Proceedings of the National Academy of Sci-
ences 109: 2418-2422.

Edwards, K. R., and C. D. Proffitt. 2003. Comparison of wetland structural charac-
teristics between created and natural salt marshes in southwest Louisiana, USA.
Wetlands 23: 344-356.

Erwin, K. L. 1991. An Evaluation of Wetland Mitigation in the South Florida Water
Manapiement District, Vol. 1. Final Report to South Florida Water Management
District, West Palm Beach, EL. 124 pp.

Gallihugh, J. L., and J. D. Rogner. 1998. Wetland Mitigation and 404 Permit Com-
pliance Study, Vol. 1. U.S. Fish and Wildlife Service, Region III, Burlington, IL,
and U.S. Environmental Protection Agency, Region V, Chicago. 161 pp.

Garbisch, E. W. 1977. Recent and Planned Marsh Establishment Work throughout the
Contiguous United States: A Survey and Basic Guidelines. GR D-77-3, U.S. Army
Gorps of Engineers Waterways Experiment Station, Vicksburg, MS.

Garbisch, E. W. 2005. Hambleton Island restoration: Environmental Goncern’s first
wetland creation project. Ecological Engineering 24: 289-307.

Garbisch, E. W, P. B. Woller, and R. J. McGallum. 1975. Salt Marsh Establishment
and Development. Technical Memo 52. U.S. Army Goastal Engineering Research
Genter, Fort Belvoir, VA.

Gorham, E., and L. Rochefort. 2003. Peatiand restoration: A brief assessment with
special reference to Sphagnum bogs. Wetlands Ecology and Management 11:
109-119.

Hart, D. D., T. E. Johnson, K. L. Bushaw-Newton, R. J. Horwitz, A. T. Bednarek,
D. F. Gharles, D. A. Kreeger, and D. J. Velinsky. 2002. Dam removal: GhaT
lenges and opportunities for ecological research and river restoration. BioScience
52: 669-681.

Haynes, R. J. 2004. The development of bottomland forest restoration in the Lower
Mississippi River alluvial valley. Ecological Restoration 22: 170-182.

Hefner, J. M., and J. D. Brown. 1985. Wetland trends in the southeastern United
States. Wetlands 4: 1-11.

Hinkle, R., and W. J. Mitsch. 2005. Salt marsh vegetation recovery at salt hay farm
wetland restoration sites on Delaware Bay. Ecological Engineering 25: 240-251.

642 Chapter 18 Wetland Creation and Restoration

Hopple, A., and C. Craft. 2013. Managed disturbance enhances biodiversity of
restored wetlands in the agricultural Midwest. Ecological Engineering 61P:
505-510.

Isselin-Nondedeu, F., L. Rochefort, and M. Poulin. 2007. Long-term vegetation
monitoring to assess the restoration success of a vacuum-mined peatland Inter-
national Conference on Peat and Peatlands 2007, 23: 153-166. Federation des
Conservatoires d’Espaces Naturels, Quebec, Canada.

Josselyn, M., J. Zedler, and T. Griswold. 1990. Wetland mitigation along the Pacific
coast of the United States. In J. A. Kusler and M. E. Kentula, eds. Wetland Cre-
ation and Restoration. Island Press, Washington, DC, pp. 3-36.

Koebel, J. W, and S. G. Bousquin. 2014. The Kissimmee River restoration project
and evaluation program, Florida, U.S.A. Restoration Ecology 22: 345-352.

Kristensen, E. A., B. Kronvang, P. Wiberg-Larsen, H. Thodsen, G. Nielsen, E. Amor,
N. Friberg, M. L. Pedersen, and A. Baattrup-Pedersen. 2014. 10 years after
the largest river restoration project in Northern Europe: Hydromorphological
changes on multiple scales in River Skjern. Ecological En^ineerin£i 66: 141-149.

Lambert,). M. 1964. The Spartina s,tocy. Nature 204-. 1136-1138.

Lewis, R. R. 1990a. Wetland restoration/creation/enhancement terminology: Sug-
gestions for standardization. In J. A. Kusler and M. E. Kentula, eds. Wetland
Creation and Restoration. Island Press, Washington, DG, pp. 1-7.

Lewis, R. R. 1990b. Greation and restoration of coastal plain wetlands in Florida. In
J. A. Kusler and M. E. Kentula, eds. Wetland Creation and Restoration. Island
Press, Washington, DG, pp. 73-101.

Lewis, R. R. 1990c. Greation and restoration of coastal plain wedands in Puerto Rico
and the U.S. Virgin Islands. In J. A. Kusler and M. E. Kentula, eds. Wetland
Creation and Restoration. Island Press, Washington, DG, pp. 103-123.

Lewis, R. R. 2005. Ecological engineering for successful management and restoration
of mangrove forests. Ecological Engineering 24: 403-418.

Lewis, R. R. 2009. Methods and criteria for successful mangrove forest restoration.
In G. M. E. Perillo, E. Wolanski, D. R. Gahoon, and M. M. Brinson, eds. Coastal
Wetlands: An Integrated Ecosystem Approach. Elsevier, Amsterdam, the Nether-
lands, pp. 787-800.

Lewis, R. R. 2010. Mangrove field of dreams: If we build it, will they come.^ Wetland
Science and Practice 27(1): 15-18.

Lewis, R. R, and R. G. Gilmore. 2007. Important considerations to achieve successful
mangrove forest restoration with optimum fish habitat. Bulletin of Marine Science
80(3): 823-837.

Lewis, R. R., and B. Brown. 2014. Ecological mangrove rehabilitation — afield man-
ual for practitioners. Version 3. Mangrove Action Project Indonesia, Blue Forests,
Ganadian International Development Agency, and OXFAM. 275 pp.

Liptak, M. A. 2000. Water column productivity, calcite precipitation, and phosphorus
dynamics in freshwater marshes. Ph.D. dissertation. The Ohio State University,
Golumbus.

Massey, B. 2000. Wetlands Engineering Manual. Ducks Unlimited, Southern
Regional Office, Hackson, MS, 16 pp.

Mitsch, W. J. 2013. Wetland creation and restoration. In S. Levin, ed.. Encyclopedia
of Biodiversity, 2nd ed. Elsevier, Amsterdam, pp. 367-383.

References 643

Mitsch, W. J. 2014. When will ecologists learn engineering and engineers learn ecol-
ogy Ecological En£iineerin£i 65: 9-14.

Mitsch, W. J., and R. F. Wilson. 1996. Improving the success of wetland creation
and restoration with know-how, time, and self-design. Ecological Applications 6:
77-83.

Mitsch, W. J., X. Wu, R. W. Nairn, R E. Weihe, N. Wang, R. Deal, and C. E.
Boucher. 1998. Creating and restoring wetlands: A whole-ecosystem experiment
in self-design. BioScience 48\ 1019-1030.

Mitsch, W. J., and S. E. Jorgensen. 2004. Ecolojjical Enjjineerinjj and Ecosystem
Restoration. John Wiley & Sons, Hoboken, NJ. 411 pp.

Mitsch, W. J., L. Zhang, C. J. Anderson, A. Altor, and M. Hernandez. 2005. Creating
riverine wetlands: Ecological succession, nutrient retention, and pulsing effects.
Ecological Enptineerinpt 25: 510-527.

Mitsch, W. J., L. Zhang, D. F. Fink, M. E. Hernandez, A. E. Altor, C. L. Tuttle
and A. M. Nahlik. 2008. Ecological engineering of floodplains. Ecohydrology &
Hydrobiolojjy 8: 139-147.

Mitsch, W. J., L. Zhang, K. C. Stefanik, A. M. Nahlik, C. J. Anderson, B. Bernal,
M. Hernandez, and K. Song. 2012. Creating wedands: Primary succession, water
quality changes, and self-design over 15 years. BioScience 62: 237-250.

Moreno-Mateos, D., M. E. Power, F. A. Comm , and R. Yockteng 2012. Structural
and functional loss in restored wedand ecosystems. PLoS Biol 10(1): el001247.
doi:10.1371 /journal. pbio. 1001247

Nahlik, A. M., and W. J. Mitsch. 2010. Methane emissions from created riverine wet-
lands. Wetlands 30: 783-793. Erratum in Wetlands {2011) 31: 449-450.

Nadonal Research Council. 2001. Compensating for Wetland Losses under the Clean
Water Act. Nadonal Academy Press, Washington, DC. 158 pp.

Nadonal Research Council. 2014. Projjress toward Restorinjj the Everjjlades: The Fifth
Biennial Review. National Academies Press, Washington, DC. 240 pp.

Odum, W. E. 1987. Predicdng ecosystem development following creation and restora-
don of wedands. In J. Zelazny and J. S. Feierabend, eds. Wetlands: Increasinp; Our
Wetland Resources. Proceedings of the Conference Wetlands: Increasinp; Our Wet-
land Resources. Corporate Conservation Council, National Wildlife Federation,
Washington, DC, pp. 67-70.

Palmer, M. A., E. S. Bernhardt, J. D. Allan, P. S., Lake, G. Alexander, S. Brooks, J.
Carr, S. Clayton, C. N. Dahm, J. Follestad Shan, D. L. Galat, S. G. Loss, P. Good-
win, D. D. Hart, B. Hassett, R. Jenkinson, G. M. Kondolf, R. Lave, L. J. Meyr,
T. K. O’Donnell, L. Pagano, and E. Sudduth. 2005. Standards for ecologically
successful river restoradon. Journal of Applied Ecology 42: 208-217.

644 Chapter 18 Wetland Creation and Restoration

Palmer, M. A., S. Filoso, and R.M. Fanelli. 2014. From ecosystems to ecosystem ser-
vices: Stream restoration as ecological engineering. Ecol. En£i. 65: 62-70.

Pedersen, M. L., J. M. Andersen, K. Nielsen, and M. Linnemann. 2007a. Restoration
of Skjern River and its valley: Project description and general ecological changes
in the project area. Ecological En£iineerin£i 30: 131-144.

Pedersen, M. L., Friberg, N., Skriver, J., Baattrup-Pedersen, A., 2007b. Restoration
of Skjern River and its valley — short-term effects on river habitats, macrophytes
and macroinvertebrates. Ecol. Eng;. 30: 145-156.

Peterson, S. B., J. M. Teal, and W. J. Mitsch, eds. 2005. Delaware Bay Salt Marsh
Restoration. Special issue of Ecolopfical Enpiineerii^g 25: 199-314.

Philipp, K. R., and R. T. Field. 2005. Ehragmites australis in Delaware Bay

salt marshes. Ecological Engineering 2S\ 275-291.

Price, J., L. Rochefort, and F. Quinty. 1998. Energy and moisture considerations on
cutover peatlands: Surface microtopography, mulch cover and Sphagnum regen-
eration. Ecological Engineering 10: 293-312.

Primavera, J. H., J. D. Savaris, B. Bajoyo, J. D. Coching, D. J. Curnick, R. Golbeque,
A. Guzman, J. Q. Henderin, R. V. Joven, R. A. Loma, and H. J. Koldewey. 2012.
Manual on Community-Based Mangrove Rehabilitation. Mangrove Manual Series
No. 1. Zoological Society of London, London, UK. viii +240 pp.

Quinty, F., and L. Rochefort. 1997. Plant reintroduction on a harvested peat bog.
In C. C. Trettin, M. F. Jurgensen, D. F. Grigal, M. R. Gale, and J. K. Jeglum,
eds. Northern Forested Wetlands: Ecology and Management . CRG Press/Lewis
Publishers, Boca Raton, FL, pp. 133-145.

Ranwell, D. S. 1967. World resources of Spartina townsendii and economic use of
marshland. Coastal Zone Management Journal 1: 65-74.

Reinartz, J. A., and E. L. Warne. 1993. Development of vegetation in small created
wetiands in southeast Wisconsin. Wetlands 153-164.

Richardson C. J., P. Reiss, N. A. Hussain, A. J. Alwash, and D. J. Pool. 2005.
The restoration potential of the Mesopotamian marshes of Iraq. Science 307:
1307-1311.

Richardson, C. J., and N. A. Hussain. 2006. Restoring the Garden of Eden: An eco-
logical assessment of the marshes of Iraq. BioScience 56: 447-489.

Robb, J. T. 2002. Assessing wetiand compensatory mitigation sites to aid in establish-
ing mitigation ratios. Wetlands 22: 435M40.

Rochefort, L., and S. Campeau. 1997. Rehabilitation work on post-harvested bogs in
south-eastern Canada. In L. Parkyn, R. E. Stoneman, and H. A. P. Ingram, eds.
Conserving Peatlands . CAB International, Walingford, UK, pp. 287-284.

Rochefort, L., G. Quinty, S. Campeau, K. Johnson, and T. Malterer. 2003. North
American approach to the restoration of Sphagnum dominated peadands.
Wetlands Ecology and Management 11: 3-20.

Rochefort, L., and E. Lode. 2006 Restoration of degraded boreal peatlands. In R.
K. Wieder and D. H. Vitt, eds. Boreal Peatland Ecosystems. Springer, Berlin,
Germany, pp. 381-424.

References 645

Rochefort, L., M. Strack, M. Poulin, J. S. Price, M. Graf, A. Desrochers, and C. Lavoie.
2012. Northern peatlands. In D. R. Batzer and A. H. Baldwin, eds. Wetland Habi-
tats of North America: Ecology and Conservation Concerns. University of California
Press, Los Angeles, pp. 119-134.

Rochefort, L., F. Isselin-Nondedeu, S. Boudreau, and M. Poulin. 2013. Comparing
survey methods for monitoring vegetation change through time in a restored
peadand. Wetlands Ecology and Management 2\-. 71-85.

Rodgers, H. L., F. P. Day, and R. Addnson. 2004. Root dynamics in restored and natu-
rally regenerated Adandc white cedar wedands. Restoration Ecolojjy 16: 401-411.

Rodgers, L., M. Bodle, D. Black, and F. Laroche. 2014. Status of nonindigenous
species. In 2014 South Elorida Environmental Report, Volume 1: The South Florida
Environment. South Florida Water Management District, West Palm Beach,
Chapter 7, pp. 7-1-7-55.

Samson, M. S., and R. N. Rollon. 2008. Growth performance of planted red man-
groves in the Philippines: Revisiting forest management strategies. Ambio 37(4):
234-240.

Sanville, W, and W. J. Mitsch, eds. 1994. Creating Freshwater Marshes in a Riparian
Landscape: Research at the Des Plaines River Wetland Demonstradon Project.
Special Issue of Ecological Engineering 3(4): 315-521.

Sartoris, J. J., J. S. Thullen, L. B. Barber, and D. E. Salas. 2000. Investigadon of nitro-
gen transformadons in a southern California constructed wastewater treatment
wedand. Ecological Engineering 14: 49-65.

Seliskar, D. 1995. Exploidng plant genotypic diversity for coastal salt marsh creadon
and restoration. In M. A. Khan and I. A. Ungar, eds. Biology of Salt-Tolerant
Plants. Department of Botany, University of Karachi, Pakistan, pp. 407-416.

Sha, C., W. J. Mitsch, U. Mander, J. Lu, J. Batson, L. Zhang and W. He. 2011.
Methane emissions from freshwater riverine wetlands. Ecological Engineering 37:
16-24.

Shin, J. Y., F. Ardgas, C. Hobble, and Y. S. Lee. 2013. Assessment of anthropogenic
influences on surface water quality in urban estuary, northern New Jersey: Muld-
variate approach. Environmental Monitoring and Assessment 185: 2777-2794.

South Florida Water Management District. 2014. South Elorida Environmental Report
2014, Executive Summary. SFWMD, West Palm Beach, FL. 28 pp.

Spieles, D. J. 2005. Vegetadon development in created, restored, and enhanced mit-
igadon wedand banks of the United States. Wetlands 25: 51-63.

Teal, J. M., and M. P. Weinstein. 2002. Ecological engineering, design, and con-
strucdon consideradons for marsh restorations in Delaware Bay, USA. Ecolopfical
Engineering 18: 607-618.

Toth, L. A., D. A. Arrington, M. A. Brady, and D. A. Muszick. 1995. Conceptual
evaluation of factors potendally affecdng restoration of habitat structure within
the channelized Kissimmee River ecosystem. Restoration Ecology 3: 160-180.

Tutde, C. L., L. Zhang, and W. J. Mitsch. 2008. Aquadc metabolism as an indicator of
the ecological effects of hydrologic pulsing in flow-through wedands. Ecological
Indicators 8: 795-806.

646 Chapter 18 Wetland Creation and Restoration

United Nations Environmental Programme. 2001. The Mesopotamian Marshlands:
Demise of an Ecosystem. UNEP/DEWA/TR.01-3 Revision 1. UNEP, Nairobi,
Kenya.

United Nations Environmental Programme. 2007. After the Tsunami — Coastal Ecosys-
tems Restoration — Lessons Learnt. UNEP, Nairobi, Kenya, iii +55 pp.

United Nations Environmental Programme. 2008. UNEP Post-Tsunami Recovery
Activities 2004— 2007. UNEP, Nairobi, Kenya, ii +68 pp.

U.S. Army Corps of Engineers and U.S. Environmental Protection Agency. 2008.
Compensatory Mitipjation for Losses of Aquatic Resources; Final Rule. Federal Reg-
ister (April 10, 2008) 73 (70): 19594-19705.

U.S. Environmental Protection Agency and U.S. Army Corps of Engineers.
2008. Wetlands Compensatory Mitiptation Rule, http://water.epa.gov/lawsregs
/guidance / wetlands / upload /MitigationRule .pdf.

Vymazal, J. 2013. Emergent plants used in free water surface constructed wetlands:
A review. Ecological Engineering 6\P\ 582-592.

Waddington, J. M., M. J. Greenwood, R. M. Petrone, and J. S. Price. 2003. Mulch
decomposition impedes recovery of net carbon sink function in a restored peat-
land. Ecolopiical Enpiineerinpi 2Q-. 199-210.

Waddington, J. M., K. Toth, and R. Bourbonniere. 2008. Dissolved organic carbon
export from a cutover and restored peatland. Hydrolopfic Processes 22: 2215-2224.

Weihe, P. E., and W. J. Mitsch. 2000. Garden wedand experiment demonstrates
genetic differences in soft rush obtained from different regions (Ohio). Ecolopfical
Restoration 18: 258-259.

Wilson, R. F., and W. J. Mitsch. 1996. Functional assessment of five wetlands con-
structed to mitigate wetland loss in Ohio, USA. Wetlands 16: 436-451.

Woodhouse, W. W. Jr. 1979. Building Salt Marshes along the Coasts of the Continental
United States. Special Report 4, U.S. Army Goastal Engineering Research Genter,
Fort Belvoir, VA.

Wu, X. and W. J. Mitsch. 1998. Spatial and temporal patterns of algae in newly con-
structed freshwater wetlands. Wetlands 18: 9-20.

Zedler, J. B. 1988. Salt marsh restoration: Lessons from Galifornia. In J. Gairns,
ed.. Rehabilitating Damaged Ecosystems, Vol. 1. GRG Press, Boca Raton, EL,
pp. 123-138.

Zedler, J. B. 2000a. Progress in wetland restoration ecology. TREE 15: 402-407.

Zedler, J. B. 2000b. Handbook for Restoring Tidal Wetlands. GRG Press, Boca Raton,
EL, 464 pp.

Chapter 19

Wetlands and Water Quality

Wetlands created specifically to improve water quality are usually referred to as
treatment wetlands. There are three types of wetlands used to treat wastewater
or stormwater: natural wetlands, surface-flow constructed wetlands, and
subsurface-flow constructed wetlands. Studies on usinp; subsurface wetlands
bepfan in the 1960s in Europe. The use of natural wetlands to treat wastewater
in Florida and Michigan in the 1970s pioneered the use of surface-flow-
created wetlands. Wetlands have been used to treat a variety of threats to
downstream water quality, includinpt domestic wastewater, mine drainage,
non-point source pollution, stormwater runoff, landfill leachate, and confined
livestock operations. The desip/n of treatment wetlands requires particular
attention to hydrology, chemical loading, soil physics and chemistry, and
wetland vepfetation. Manapfement issues include wildlife control and
attraction, mosquito and pathogen control, and pfreenhouse pfas and water-level
manapfement. Treatment wetlands are not inexpensive to build and operate,
but they usually cost much less than chemical and physical treatment systems.

Wastewater and polluted water treatment by wetlands is an intriguing concept involv-
ing the forging of a partnership between humanity (our wastes) and an ecosystem
(wetlands). Therefore, it is a good example of ecological engineering (see Mitsch and
Jorgensen, 2004, and Mitsch, 2012, for definitions and applications of this field). In
this chapter, we discuss the use of wetlands for removing unwanted chemicals from
waters, be the waters municipal wastewater, non-point source runoff, or other forms
of pollution.

As described in Chapter 6: “Wetland Biogeochemistry,” wetlands can be sources,
sinks, or transformers for a great number of chemicals. A wetland is a sink if it has
a net retention of an element or a specific form of that element (e.g., organic or

647

648 Chapter 19 Wetlands and Water Quality

inorganic); that is, if the inputs are greater than the outputs. The desired situation
for treatment wetlands is to optimize the wetlands’ ability to serve as chemical (and
sometimes biological) sinks.

German scientists investigated the use of constructed basins with macrophytes
{hohere Pflanzen) for purification of wastewater. Later, researchers in Florida and
Michigan in the United States investigated the role of natural wetlands to treat
wastewater and thus recycle clean water back to groundwater and surface water. The
two different approaches, one using artificial systems and the other utilizing natural
wetlands, have converged into the general field of treatment wetlands (Kadlec and
Wallace, 2009). The field now encompasses the construction and/or use of wedands
for a myriad of water quality applications. While water quality improvement is the
primary goal of treatment wetlands, they also provide habitat for a wide diversity of
plants and animals and can support many of the other wetland functions and services
described in this book.

Classifications of Wastewater Treatment Wetlands

Three General Approaches

Three types of wetlands are used to treat wastewater. In the first approach, wastewater
is purposefully introduced to existing natural wetlands rather than constructed wet-
lands (Fig. 19.1a). In the 1970s, studies involving application of wastewater to natural
wetlands were carried out in locations of the United States such as Michigan (Kadlec,
2009b) and Florida (Odum et al., 1977; Ewel and Odum, 1984) where there were
abundant wetlands. At that time, legal protection of wetlands had not been institu-
tionalized. These pioneering studies elevated the importance of wetlands as “nature’s
kidneys” to the general public and governmental agencies. This importance was then
translated, appropriately, into laws that protected wetlands. These same laws now often
prohibit the addition of wastewater or polluted water to natural wetlands.

Two types of constructed wetlands are alternatives to using natural wetlands.
Surface-flow constructed wetlands (Fig. 19.1b) mimic natural wetlands and can be a
better habitat for certain wetland species because of standing water through most
if not all of the year. Subsurface-flow constructed wetlands (Fig. 19.1c) more closely
resemble wastewater treatment plants than wetlands. In these systems, the water
flows horizontally or vertically through a porous medium, usually sand or gravel,
supporting one or two of a relatively narrow list of macrophytes, such as Phra£imites
australis. There is rarely standing water in these subsurface systems as the wastewater
passes laterally through the medium.

Subsurface treatment wetlands had their start in the Max-Planck Institute in
Germany in the 1950s. Dr. Kathe Seidel performed many experiments with emergent
macrophytes, Schoenoplectus lacustris in particular, and found that the plants con-
tributed to the reduction of bacteria, and organic and inorganic chemicals (Seidel,
1964, 1966). This process was translated into a gravel-bed macrophyte system that
became known as the Max-Planck Institute process or the Krefeld system (Seidel and
Happl, 1981, as cited in Brix, 1994). The development of subsurface wetlands

Classifications of Wastewater Treatment Wetlands 649

Natural Wetland

outlet weir

Surface-Flow Constructed Wetland

distribution pipe

f '

' N
f

f ’

! '
'

, a(

jjustable
and pipe

f i-

} ? / / y

f / f ?/////// /

Gravel or soil matrix

Subsurface-Flow Constructed Wetland

Figure 19.1 Three types ef wetland treatment systems: (a) natural wetland, (b) surface-flow
wetland, and (c) subsurface-flow wetland (After Kadlec and Knight, 1996 and Kadiec and
Waiiace, 2008)

continued in Europe using a system of subsurface-flow basins planted with P. australis.
These systems were called the root-zone method {Wurzelraumentsor£iun£i) . Subsurface
wetland systems continued to be studied and refined through the work of Dejong
(1976) in Holland, Brix (1987) in Denmark, and many other scientists in Europe. The
appeal of these more “artificial” types of wetlands in Europe (as opposed to free-water
surface wetlands in North America) is due to two factors: (1) there are fewer natural
wetlands remaining in Europe, and those that are left are protected for nature; and

650 Chapter 19 Wetlands and Water Quality

(2) space is much more at a premium in Europe, and subsurface wedands require less
land area.

Classification According to Vegetation

Treatment wedands can also be classified based on the life-form of their vegetation.
In this case, there are five systems based on their vegetation:

1. Free-floating macrophyte systems — for example, water hyacinth {Eichhornia
crassipes), duckweed {Lemna- spp.)

2. Emergent macrophyte systems — for example, reed grass {P. a-ustmlis), cattails
( Typha spp.)

3. Submerged aquadc vegetation systems

4. Forested wedand systems

5. Muldspecies algal systems, pardcularly algal-scrubber systems

Subsurface-flow constructed wetlands are limited to emergent macrophytes,
whereas surface-flow constructed wetlands often utilize a combinadon of free-floadng,
emergent, and submerged macrophytes. Forested wedand treatment systems are
generally not constructed wedands at all but are natural wedands to which wastewater
is applied. They will often develop extensive communides of all of the other vegetadon
types described in this classificadon.

Treatment Wetland Types

The type of wastewater being treated can classify treatment wedands. While many
of these systems are used for municipal wastewater and that is often thought as the
convendonal system, there has been much interest in the use of wetlands to treat
stormwater from urban areas, acid mine drainage from coal mines, non-point source
polludon in rural landscapes, livestock and aquaculture wastewaters, and an array of
industrial wastewaters.

Municipal Wastewater Wetlands

In Europe, most of the development of subsurface constructed wedands was to
replace both primary and secondary treatment to remove biochemical oxygen demand
(BOD) and suspended solids as well as inorganic nutrients. Hundreds of subsurface
wetland treatment systems for municipal wastewater have been constructed in
Europe, pardcularly in the United Kingdom (Cooper and Findlater, 1990), Denmark
(Brix and Schierup, 1989a, b; Brix, 1998; Brix and Arias, 2005), the Czech Republic
(Vymazal, 1995, 1998, 2002, 2005; Vymazal and Kropfelova, 2005), Norway
(Braskerud, 2002a, b), Spain (Solano et ah, 2004), and Estonia (Teiter and Mander,

2005) . There are also many applicadons of this technology in Australia (Mitchell et ah,
1995; Greenway et ah, 2003; Greenway, 2005; Headley et al., 2005; Davison et ah,

2006) , New Zealand (Gooke, 1992; Tanner, 1996; Nguyen et al., 1997; Nguyen,

Classifications of Wastewater Treatment Wetlands 651

2000), and Costa Rica (Nahlik and Mitsch, 2006). In North America, most but
certainly not all of the wetlands built for treatment of municipal wastewater treatment
are surface-water wetlands. Locations of wastewater wetlands that have been studied
in some detail include Florida (Knight et ah, 1987; J. Jackson, 1989), California
(Gerheart et al., 1989; Gerheart, 1992; Sartoris et al., 2000; Thullen et ah, 2005),
Louisiana (Boustany et ah, 1997; Day et ah, 2004), Arizona (Wilhelm et ah, 1989),
Kentucky (Steiner et ah, 1987; Steiner and Freeman, 1989), Pennsylvania (Conway
and Murtha, 1989), Ohio (Spieles and Mitsch, 2000a, b). North Dakota (Litchfield
and Schatz, 1989), and Alberta, Canada (White et ah, 2000). Created wetlands for
treating wastewater have been most effective for controlling organic matter (BOD),
suspended sediments, and nutrients. Their value for controlling trace metals and other
toxic materials is more controversial, not because these chemicals are not retained in
the wetlands but because of concerns that they might concentrate in wetland substrate
and fauna.

One of the longest running treatment wetland as far as data collection is the
Houghton Lake treatment wetland in Michigan that had its beginnings in the early
1970s. This is described in detail in Case Study 1.

CASE STUDY 1: Long-Term Effectiveness of a Treatment Wetland at
Houghton Lake, Michigan

Much of the interest in using surface-flow wetlands for water quality manage-
ment was sparked by several studies begun in the early 1970s. In one of those
studies, peatlands in Michigan were investigated by researchers from the Uni-
versity of Michigan for the wetlands’ capacity to treat wastewater (Fig. 19.2).
A pilot operation for disposing of up to SSOm^day (100,000 gallons per
day) of secondarily treated wastewater in a 700-ha rich fen at Houghton Lake
led to significant reductions in ammonia nitrogen and total dissolved phos-
phorus as the wastewater passed from the point of discharge through the
wetlands. Inert materials, such as chloride, did not change as the wastewater
passed through the wetland. In 1978, the flow was increased to approximately
5,000 m^ day^ over a much larger area, essentially using all the wastewater
from the local treatment plant. Data after 30 years of operation show that the
wetlands have remained effective over that long time in removing both total
phosphorus (Fig. 19.2b,c) and inorganic nitrogen (Fig. 19. 2d). Approximately
100 ha of the peatland was an irrigation zone (Fig. 19.2a) where most of the
water quality improvement occurred (Kadlec, 2009b). Phosphorus decreased
by 94 percent from 3.5 to 0.04 ppm (= mg-P/L) and dissolved inorganic nitro-
gen decreased by 95 percent. Nutrient retention on an annualized basis was
1.76 g-P m“^ yr“^ for phosphorus and 4.39 g-N m“^ yr“^ for dissolved inor-
ganic nitrogen (Kadlec, 2009b). The peatland itself was not unaffected, with
the vegetation in the irrigation area changing from a sedge-Sa//x community to

652 Chapter 19 Wetlands and Water Quality

a Typha spp. community, some of which has become floating mats (Kadlec,
2009b).

0.1

^ 0.01

0.001

■

■ a ■ a

■

^ .

Pumped HLSA <>

A •

Pumped WMS s

Outlet HLSA a

Outlet WMS o 1

1 1

0 5

10 15

20 25 30

Figure 19.2 Houghton Lake treatment wetland in Michigan, where treated wastewater
has been discharged into a peatland for 30 years: (a) map of site showing the irrigation
area; (b) total phosphorus of influent and outlet as measured by Houghton Lake Sewer
Authority (HLSA) and Wetland Management Services (WMS); (c) phosphorus inflow
and outflow fluxes; (d) dissolved inorganic nitrogen inflow and outflow fluxes. (After
Kadlec, 2009b)

Classifications of Wastewater Treatment Wetlands 653

0 5 10 15 20 25 30

Years of Operation

Figure 19.2 {Continued)

Mine Drainage Wetiands

Wetlands frequently have been used as downstream treatment systems for mineral
mines. An example of such a system in southeast Ohio is described in Figure 19.3.
Acid mine drainage water, with its low pH and high concentrations of iron, sulfate,
aluminum, and trace metals, is a major water pollution problem in many coal min-
ing regions of the world, and constructed wetlands are a viable treatment option.
The use of wetlands for coal mine drainage control was probably first considered
when volunteer Typha wetlands were observed near acid seeps in a harsh environ-
ment where no other vegetation could grow. By the 1980s, hundreds of wedands
had been constructed in the eastern United States alone to treat mine drainage water.
The most common goal of these systems was usually the removal of iron from the
water column to avoid its discharge downstream, but sulfate reduction and the alle-
viation of extremely acidic conditions were also appropriate goals (Wieder and Lang,
1984; Brodie et al., 1988; Fennessy and Mitsch, 1989; Mitsch and Wise, 1998; Tarutis
et al., 1999).

654 Chapter 19 Wetlands and Water Quality

b. 350
300
^ 250

c-200

S150

100

Figure 19.3 (a) A 0.4-ha acid mine drainage treatment wetiand in southeastern Ohio; (b)
totai iron in wetland and downstream before wetland was built and after it was constructed.
(After Mitsch and Wise, 1998)

Design criteria for these wetlands have been developed, but they are neither con-
sistent from site to site nor generally accepted. Some are suggested in Table 19.1 . Stark
and Williams (1995) found design features that enhanced iron removal and decreased
acidity included broad drainage basins, nonchannelized flow patterns, high plant diver-
sity, southern exposure, low flow rates and loadings, and shallow depths. It is not

Classifications of Wastewater Treatment Wetlands 655

Table 19.1 Suggested design parameters for constructed wetlands used for controlling coal
mine drainage

Parameter

Design

Reference

Hydrologic loading rate, cm/day

Fennessy and Mitsch, 1989

Retention time, days

Fennessy and Mitsch, 1989

Iron loading, g-Fe day“^

pH <5.5

0.72

Brodie et at, 1988

pH >5.5

1.29

Brodie et at, 1988

For 90% removal, pH =6

2-10

Fennessy and Mitsch, 1989

For 50% removal, pH =6

20-40

Fennessy and Mitsch, 1989

pH =3.5, onflow <3.5 mg-Fe/L

2.5

Manyin et al., 1997

Basin characteristics

Depth, m

<0.3

Number of cells

Plant material

Typha spp.

Substrate material

Organic peat over clay seal;

spent mushroom material

always cost effective to construct wetlands when extremely high (>85-90 percent)
iron removal efficiencies are necessary or when the pH of the mine drainage water
is less than 4. Hydraulic loading rates as high as 29 cm day”'^ have been suggested
for wetlands designed for acid mine drainage, although Fennessy and Mitsch (1989)
recommended 5 cm/day as a conservative loading rate for this type of wetland and a
minimum detention time of 1 day, with much longer periods for more effective iron
removal (Table 19.1). The long-term suitability of wetland treatment systems is poorly
understood, although it appears that Typha-domm-atcd systems can survive decades in
a mine drainage system. The accumulation of iron hydroxides can eventually cause
mine drainage systems to begin to export materials, unless the design and manage-
ment includes adequate storage capacity and/or material removal. Some researchers
suggest that these wetlands, after several decades, can become mineral mines in their
own sense, effectively recycling minerals that otherwise would be lost to downstream
watersheds back to the economy. Where no other alternative is feasible, the use of wet-
lands to reduce this harsh water pollution should be viewed as a reasonable, low-cost
alternative.

Urban Stormwater Treatment Wetlands

The control of stormwater pollution with wetlands is a valid and increasingly utilized
application of wetland ecological engineering. Unlike municipal wastewater, stormwa-
ter and other non-point source pollution are seasonal, often quite sporadic, and vari-
able in quality, depending on season and recent land use. Wedands are one of several
choices for systems to control urban runoff More conventional approaches involve
either dry detention ponds that fill only during storms or wet detention ponds that
are usually deepwater systems, where the edge is usually stabilized with rocks and plant
growth is actually discouraged.

656 Chapter 19 Wetlands and Water Quality

Figure 19.4 General design of a stormwater treatment wetland. (After Schueler, 1992)

Stormwater from urban areas is particularly rapid as it comes from impervious
sources such as roofs, parking lots, and highways. One of the features of stormwater
wetland systems is that severe storms have a dramatic effect on treatment efficiency.
High flows resulting from high-intensity rainstorms usually result in lower nutrient
and other chemical retention as a percentage of inflow, and sometimes the storms
cause a net release of nutrients. The very nature of the sudden but short stormwater
pulses makes management of these systems particularly difficult.

A layout of an ideal stormwater treatment wetland (Fig. 19.4) illustrates that a
combination of deep ponds and marshes may be most appropriate. The first “wet
pond” is a deep, usually unvegetated, basin designed to dampen the rapid stormwater
pulse, allowing the downstream marshes to “treat” the runoff in a more effective
manner. Multiple cells of marshes and a small outflow deepwater pond can contribute
to the system’s effectiveness. Sediment retention capability is the strong point of these
wetlands, but if any significant construction projects occur upstream, even this capacity
can be temporarily or permanently overwhelmed. Case Study 2 presents an application
of this design.

CASE STUDY 2: Freedom Park: Intercepting Urban Runoff in an Urban
Wetiand Park in Fiorida

A 20-ha constructed wetland complex, called Freedom Park, was constructed
in 2007-2008 in Naples, Florida, at an abandoned citrus grove to treat urban
stormwater runoff (Fig. 19.5a). Its total construction cost was $10 million,
and it was designed to treat 757,000 m^ per year for an average hydraulic
loading rate (HLR) of about 7.6m/yr (2cm/day). The system was designed

(b)

Figure 19.5 Freedom Park stormwater treatment wetlands in Naples Florida: (a) photo
looking west with restored upland and wetland forests in the foreground and stormwa-
ter treatment wetland basins in the background; (b) map of Freedom Park, illustrating
stormwater treatment ponds and wetlands on the left and restored forested wetlands
on the right. Also shown are boardwalks through the system and visitors center;

(d) total phosphorus in the inflow and outflow of the stormwater wetlands. (Figures a,
c, and d courtesy of Jim Bays, CH2M-Hill, Tampa, FL)

657

658 Chapter 19 Wetlands and Water Quality

♦ Stormwater inftow cone. □ wetland outftow cone.

• ■ • stormwater inflow median — ■ wetland outflow median
O Gordon River concentration

Figure 19.5 {Continued)

especially to treat urban stormwater runoff during the summer wet season
when average flow is 10 to 100 times higher than in the dry season. During that
wet season, the design calls for an average water detention time of 18 days.
The wetland system includes an initial 1.9-ha deepwater pond that receives
and temporarily stores the stormwater pulses, followed by 2.7 ha of shal-
lower vegetated wetland ponds designed to sequentially improve water quality
(Fig. 19.5b). The wetlands were planted with Nuphar, Nymphaea, Thalia, Pont-
ederia, Eleocharis, Sagittaria, Cladium, and submerged aquatic vegetation.
The last basin was designed as a shallow limestone bed vegetated with a
periphyton-£/eochar/s community for enhanced phosphorus removal. The water
from the ponds/wetlands is then discharged as sheet flow through about 6 ha
of restored forested hardwood hammock/wetland area before it flows into to
the Gordon River. Water quality results from 2008 to 2012 show an average
decrease of 37 percent of total nitrogen (Fig. 19.5c) and a decrease of total

Classifications of Wastewater Treatment Wetlands 659

phosphorus of 81 percent (Fig. 19. d) through the ponds. Retention of both
nutrients is important as the Gordon River flows first to Naples Bay in “old
Naples” to the south and then to the coastal waters of the Gulf of Mexico.
The treatment wetland site is also a city/county park with a visitors’ center,
several kilometers of 2-m-wide boardwalks, and many interpretative signs. An
estimated 25,000 visitors came to Freedom Park in 2013 (J. Bays, personal
communication).

Agricultural Stormwater Treatment Wetlands

One of the most important applications of wetland treatment systems — ^yet an
application that is still behind municipal treatment wetlands in understanding design
issues — ^is the use of non-point source wetlands for treating stormwater and runoff
from agricultural fields. Research projects illustrating the effects and functioning of
these types of wedands in agricultural watersheds have been carried out in south-
eastern Australia (Raisin and Mitchell, 1995; Raisin et al., 1997), northeastern Spain
(Comin et al., 1997), Illinois (Kadlec and Hey, 1994; Phipps and Crumpton, 1994;
Mitsch et al., 1995; Kovacic et al., 2000; Larson et al., 2000; Hoagland et al., 2001),
Florida (Moustafa, 1999; Reddy et al., 2006), Ohio (Fink and Mitsch, 2004), and
Sweden (Leonardson et al., 1994; Jacks et al., 1994; Arheimer and Wittgren, 1994).
Several wedand sites have received the equivalent of non-point source polludon but
under somewhat controlled hydrologic condidons (e.g., river overflow to riparian
basins) over several years of study. Bony Marsh, a constructed wetland located along
the Kissimmee River in southern Florida, was investigated for nutrient retendon of
river water for nine years (1978-1986) by Moustafa et al. (1996), who found it to be
a consistent sink of nitrogen and phosphorus but at reladvely low levels.

As described in Chapters 14: “Human Impacts and Management of Wetlands,”
and 18: “Wedand Creadon and Restoration,” the water quality in the Florida Ever-
glades is threatened by high nutrients coming from the upstream Everglades Agri-
cultural Area (EAA). Case Study 3 describes what is probably the largest assemblage
of treatment wedands anywhere in the world. They have been created to intercept
phosphorus coming from agricultural stormwater runoff before it reaches the Florida
Everglades.

CASE STUDY 3: Creating Treatment Wetlands to Protect Downstream
Wetlands in Florida’s Everglades

An ambitious plan of agricultural stormwater treatment by wetlands is occur-
ring in the Florida Everglades, where 23,000 ha of created wetlands, called
stormwater treatment areas (STAs), have been created for phosphorus control
from upstream agricultural areas. As described in Chapter 14, the main cause
of the spread of cattail {Typha domingensis) in the otherwise nutrient-poor

660 Chapter 19 Wetlands and Water Quality

Florida Everglades dominated by sawgrass {Cladium jamaicense) is nutrient
enrichment — especially by phosphorus emanating from agricultural areas in
the basin.

A prototype of the STAs, a 1,544-ha treatment wetland complex called the
Everglades Nutrient Removal (ENR) project, was first designed and tested, and
the results were published in a special issue of Ecological Engineering (Reddy
et al., 2006). Water was pumped to the ENR wetland complex from adjacent
drainage canals. Over its first six-year operating schedule (1994-1999), the
wetland decreased total phosphorus and total nitrogen by 79 and 26 percent,
respectively (Gu et al., 2006), with an average outflow concentration of 21
ppb (= [ig-P/L over that period (Kadlec, 2006).

As a result of the success of the ENR project, six full-scale STAs treating
agricultural runoff from the EAA south of Lake Okeechobee have since been
created (Fig. 19.6a). Some of these systems have now been in operation for
almost 20 years (Fig. 19.6b). Overall, from their start through 2012, these
wetlands reduced phosphorus loads by 73 percent and lowered the average
phosphorus concentrations from 140 to 37 ppb (Pietro, 2012; Fig. 19.6c).
The dynamics of phosphorus in these STA wetlands has been investigated
by a number of researchers (Newman and Pietro, 2001; Juston and DeBusk,
2006, 2011; Dierberg and DeBusk, 2008; Paudel et al., 2010; Paudel and
Jawitz, 2012; Entry and Gottlieb, 2014). Juston and DeBusk (2006) suggested
that mass loadings at or below 1.3 g P m"^ yr^ provided “a high likelihood of
achieving outflow total P (TP) concentrations less than 30iig/L." Submerged
aquatic vegetation wetlands and emergent vegetation wetlands restored from
historic wetlands rather than agriculture and with loading rates at or below 2 g
P m~2 yr^ have resulted in outflow phosphorus concentrations consistently
between 10 and 20 ppb and mass removal efficiencies consistently above
85 percent. For the period 2008 to 2012, the inflow concentration of phos-
phorus to the STAs was 191 ppb and an average outflow concentration was
35 ppb, resulting in an average retention rate, by concentration, of 82 per-
cent. The average retention rate of 1.25 g-P yr^ is well within the range
of phosphorus retention of 0.5 to 5 g-P rr\~^ yr^ for low-nutrient (nonmunici-
pal wastewater) “sustainable" treatment wetlands as summarized by Mitsch
et al. (2000). These STA wetlands were originally designed to reduce phospho-
rus to 50 ppb, a 60 to 75 percent reduction in concentration (Chimney and
Goforth, 2006).

That effluent goal has been reduced by authorities to 10 ppb, essentially
the background concentrations of phosphorus in the oligotrophic Everglades.
Reaching this mandated 10 ppb threshold of total phosphorus has not been
achieved with any consistency from the STAs. A multiyear mesocosm study
that investigated the effect that different plant communities had on reducing

Classifications of Wastewater Treatment Wetlands 661

the outflow of the STAs to 10 ppb of phosphorus showed that 10 ppb may be
possible when the Inflow Is the effluent coming from the STAs and the hydraulic
loading rates are substantially lower than those in the current STAs (Mitsch
etal., 2015).

Figure 19.6 Stormwater treatment areas (STAs) downstream of the Everglades Agri-
cultural Area created to reduce phosphorus inflow to the Florida Everglades: (a) loca-
tion of the STAs; (b) inflow and outflow concentrations of total phosphorus in STA IW,
1995-2012; (c) comparison of phosphorus concentrations in inflows and outflows of
six STA complexes. (Data courtesy of South Florida Water Management District, West
Palm Beach, FL)

662 Chapter 19 Wetlands and Water Quality

i/i"

1/1

■ INFLOW

■ OUTFLOW

1E 1W 2 3/4 5

STORMWATER TREATMENT AREA

Figure 19.6 {Continued)

Agricultural Wastewater Wetlands

In addition to the nonpoint sources from agriculture that were discussed above, serious
water pollution problems occur in many parts of the world resulting from runoff from
confined animals, particularly dairy, catde, and swine operations (Tanner et ah, 1995;

Classifications of Wastewater Treatment Wetlands 663

Table 19.2 Hydrology and water quality of two wetlands constructed to
deal with heaviiy poiiuted dairy milkhouse effluent

Connecticut^ Maryland^

Wetland area,

400

1,160

Flow, m^/wk

18.8

—

retention time, days

—

Inflow

Outflow

Inflow

Outflow

BOD, mg/L

2,680

611

1,914

Total N

103

170

Ammonla-N

NItrate-N

0.3

0.1

5.5

10.0

Total F? mg/L

2.2

TSS, mg/L

1,284

130

1,645

conform, #/100 mL

557,000

13,700

—

^Newman et al., 2000.
^Schaafsma et al., 2000.

Cronk, 1996; Knight et al., 2000). As more animals are concentrated per unit area to
increase food production, the concentrations and volumes of effluents are becoming
more noticeable, both by the public and by water pollution control authorities. Con-
centrations of organic matter, organic nitrogen, ammonia-nitrogen, phosphorus, and
fecal coliforms from animal feedlots far exceed concentrations in most municipal sewer
systems. Two examples from the eastern United States of the effectiveness of wedands
for treating wastewater from dairy milkhouses (Table 19.2) showed significant reduc-
tions in most pollutants in the treated water, although ammonia-nitrogen increased
substantially in the Connecticut wetland and nitrate-nitrogen increased by 80 percent
in the Maryland case. In addition to livestock waste from land-based agriculture, con-
structed wetlands have been used to treat effluent from several aquaculture operations,
including shrimp ponds in Thailand and tilapia fish ponds in the United Kingdom.

River Diversion Wetiands

A somewhat different approach to cleaning up water is to pass river water through
wetlands built on adjacent floodplains or backwaters. These are analogs of riverine
oxbows or billabongs found throughout the world, and they have been shown to
consistently improve water quality. These wetlands also are simulations of agriculture
runoff wetlands, but with usually lower concentrations of nutrients. However, river
sediment concentrations can be high, sometimes in excess of that found in agricultural
runoff River diversions have been done on a large scale in the Mississippi River Delta
in Louisiana (see Case Study 4) and on a much smaller scale for research and water
quality improvement in the midwestern United States (see Case Study 5). In both of
these cases, significant improvement in diverted river water quality has been observed
as the water is distributed to wetlands on the floodplains and deltas.

664 Chapter 19 Wetlands and Water Quality

CASE STUDY 4: Diverting the Mississippi River to the Louisiana Deita

In the Louisiana delta, controlled diversions of the Mississippi River to the
delta to renourish the delta and slow down wetland vegetation loss have been
implemented at several sites along the river. It has also been recognized for
some time that restoration of the delta could lead to reductions of the nutri-
ent load, particularly from nitrogen, the primary cause of the Gulf of Mexico
hypoxia (see Chapter 6: “Wetland Biogeochemistry”). It has been estimated
that freshwater marshes at river diversions could denitrify up to 110 t-N km”^
yr^ and overall remove up to 25 percent of the annual flux of nitrate-nitrogen
to the Gulf, or 956,000t-N yr^ (Rivera-Monroy et al., 2013). The U.S. Army
Corps of Engineers has been and will continue to invest hundreds of millions
of U.S. dollars into river diversion systems in the Louisiana delta, now as part
of a 2012 Coastal Management Plan. One of the largest diversions in opera-
tion on the river aimed at restoring deteriorating wetlands in the Mississippi
delta is at Caernarvon (Fig. 19.7) on the east bank of the river south of New
Orleans. It has a maximum flow of 226 m^ sec~^ but an average discharge of
only 21 m^ sec~^ (Lane et al., 2006, 2007). River diversion began in August
1991, and peak discharge to date has been 140 m^ sec~^, which occurred
in March 2007 (Day et al., 2013). Summer diversion flow rates are generally
near the minimum, and winter flow rates are 50 to 80 percent of the maximum
(Lane et al., 2004). The diversion delivers river water to the 260-km^ Caernar-
von freshwater wetlands, which eventually discharge into the larger brackish
Breton Sound estuary on the Gulf of Mexico.

An intriguing issue is whether these downstream wetlands also retain
nutrients, which is particularly important given the hypoxia in the Gulf of Mex-
ico, discussed elsewhere in this book (see Chapter 6). The Caernarvon wet-
land and especially downstream Breton Sound were shown to have a loading
rate of 3.5 g-N m”^ yr^, with overall retention efficiencies of that rate high-
est in the fall (98 percent) and lowest in the winter (74 percent) (Lundberg
etal., 2014).

But as studies such as those by Deegan et al. (2012) in a whole-
ecosystem, multiyear experiment in Massachusetts salt marshes illustrate,
care must be taken so that the advantages of increasing the land elevation in
the delta to restore the marshes and retaining nutrients to protect downstream
deepwater coastal ecosystems is not outweighed by the negative effects that
excess nutrients have on salt marsh stability. There continues to be discus-
sion of the long-term effectiveness of these diversions. Some studies have
suggested that the high-nutrient inputs to the delta weaken salt marsh plant
structure and causes shallow rooting, which in turn has led to loss of vege-
tation cover during hurricanes (Kearney et al., 2011; Teal et al., 2012). This
assumption of marsh plant demise was challenged by Day et al. (2013), who
investigated wetland vegetation productivity downstream of the Caernarvon

Classifications of Wastewater Treatment Wetlands 665

Figure 19.7 Caernarvon diversion from the Mississippi River immediateiy downstream
of New Orieans and downstream Breton Sound in the Louisiana Delta. The shaded area
indicates the area of highest sedimentation in which nitrate-nitrogen concentrations
decreased by an average of 55 percent. Overall, inorganic nitrogen retention ranged
from 79 percent (fall) to 98 percent in Breton Sound estuary. (From Mitsch et al.
2005b)

river diversion in 2006-2007 following the 2005 Hurricane Katrina. They
found lower above-ground biomass than measurements before the hurricane,
but above-ground net primary productivity was near the norms of comparable
marshes with a range from 329 to 1,265 g m“^ yr^ and average 840 g m“^
yr^). Most important, below-ground alive biomass was quite high (up to
17.9 kg/m^ near the diversion structure; average 11.2 kg/m^), and sediment
accretion remained at an average Icm/yr. Morris et al. (2013) concluded the
following in a review of this debate:

Knowledge of the effect of nitrate on anaerobic soils is incomplete, but the

balance of all evidence, including studies of existing diversions and

666 Chapter 19 Wetlands and Water Quality

long-term experimental studies of sediment accretion in fertilized plots,
supports the efficacy of diverting water and sediment from the Mississippi
River to restore and stabilize its wetlands. The need for action to restore the
wetlands is urgent, and with a thoughtfully designed monitoring scheme in
place, plans to divert sediment laden water into the wetlands should
proceed.

CASE STUDY 5: Kidney-Shaped Riverine Wetiands Act Like Landscape
Kidneys

In the midwestern United States, created riparian wetlands first at the Des
Plaines River Wetlands in northeastern Illinois (Kadlec and Hey, 1994; Phipps
and Crumpton, 1994; Mitsch et al., 1995) and later at the Olentangy River
Wetland Research Park at the Ohio State University in central Ohio (Mitsch
et al., 1998, 2005a, c, 2008, 2012, 2014; Fink and Mitsch, 2007) have
shown patterns of nutrient and sediment retention over multiple years of

Figure 19.8 Olentangy River Wetland Research Park: (a) photo of pumped and nat-
urally flooded river diversion wetlands in central Ohio at the. The kidney-shaped
wetland basins in the center of the photo were constructed on a floodplain of the
Olentangy River in 1993-1994 and received pumped water from the river from March
1994 through December 2010 according to a formula of pumping rates proportional to
the river flow (Mitsch et al., 2012). Nutrient retention results for the two experimen-
tal wetlands, presented as percent change in concentrations from 1994 through 2010
are shown for (b) total phosphorus, (c) soluble reactive phosphorus, and (d) nitrate H-
nitrite nitrogen. Each data point represents the average annual decrease in concentra-
tions from inflow to outflow based on weekly sampling. (Photo from W.J. Mitsch; data
are updated from Mitsch et al., 2012).

Classifications of Wastewater Treatment Wetlands 667

b) 20

1994 1996 1998 2000 2002 2004 2006 2008 2010

Figure 19.8 {Continued)

Study. Both wetland sites received pumped and overflow river floods, thus
simulating oxbow wetlands receiving dilute non-point source pollution. For
17 years (1994-2010), the kidney-shaped experimental wetlands In Ohio
(Fig. 19.8a) consistently reduced total phosphorus, soluble reactive phos-
phorus, and nitrate -I- nitrite-nitrogen concentrations about 20 to 60 percent
(Fig. 19.8b, c, d). Both total phosphorus and soluble reactive phosphorus
showed trends of decreased retention over that 17-year period (Mitsch et al.,
2012), with the wetlands actually exporting total phosphorus in one year
(2003). Nitrate-nitrogen retention showed a steady pattern over the last
six years of the study. There has been little difference in nutrient retention
between the two experimental wetlands since they were created in 1994, even

668 Chapter 19 Wetlands and Water Quality

though one of the wetland basins was planted In 1994 (see Case Study 2
in Chapter 7: “Wetland Vegetation and Succession" and Case Study 8 in
Chapter 18: “Wetland Creation and Restoration") and the other was allowed
to colonize naturally. MItsch et al. (2014) Investigated nutrient fluxes in these
wetlands for the entire 17-year period and found that the planted wetland had
higher phosphorus retention and lower nitrogen retention. Investigation of
the most recent data showed that the trends of nutrient retention appeared
to reverse themselves for the last few years and the wetlands were actually
improving in nutrient retention.

Landfill Leachate Wetlands

Impermeable liners are used to collect groundwater that has passed through the land-
fill. This leachate is often quite variable in water quality but generally has very high
concentrations of ammonium-nitrogen and chemical oxygen demand (Kadlec, 1999).
This wastewater has always presented a problem to landfill operators, and stricter water
quality standards are making it necessary for advanced treatment. Wedands are one of
several options for management of leachate; other options include spray irrigation,
physical/chemical treatment, biological treatment, and piping to a wastewater treat-
ment plant. Mulamoottil et al. (1999) presented a summary of results from several
dozen constructed wetlands that are treating landfill leachate in Canada, the United
States, and Europe.

Water Quality Wetland Design

The need for rigor in designing a wetland varies widely depending on the site and
application. In general, a design that uses natural processes to achieve the objectives
yields a less expensive and more satisfactory solution in the long run. However, “nat-
urally” designed wetlands may not develop as predictably as more tightly designed
systems should. The choice of design is strongly affected by the site and the objec-
tives. In Europe and many parts of North America, subsurface wetlands are designed
in rectangular basins to very specific design criteria. In coastal Louisiana, by contrast,
there are now several projects where wetlands are being used as tertiary treatment sys-
tems for the removal of nutrients from wastewater. In the following sections, we focus
on rigidly designed wetlands, in part because this kind of wetland creation requires
much greater ecotechnological sophistication.

Hydrology

Hydrology is an important variable in any wetland design. If the proper hydrologic
conditions are developed, chemical and biological conditions will respond accord-
ingly. Improper hydrology leads to the failure of many created wetlands because it
will not always correct itself, as will the more forgiving biological components of

Water Quality Wetland Design 669

the system. Ultimately, hydrologic conditions determine wetland function. Several
parameters used to describe the hydrologic conditions of treatment wetlands include
hydroperiod, depth, seasonal pulses, hydraulic loading rates, and retention time.

Hydroperiod and Depth

In wetlands, hydroperiod is the water depth or stage of a wetland over time
(see Chapter 4: “Wetland Hydrology”). Wetlands that have a seasonal fluctuation of
water depth have the most potential for developing a diversity of plants, animals, and
biogeochemical processes. In a constructed wastewater wetland with a similar inflow
of wastewater every day, water levels often vary little seasonally unless stormwater is
part of the treatment inflow. During the start-up period of constructed wetlands, low
water levels are needed to avoid flooding newly emerged plants. Start-up periods for
the establishment of vegetation may take two to three years of careful attention to
water levels.

While storms and seasonal patterns of floods rarely affect constructed wastewa-
ter wetlands built for municipal treatment (except when storm sewers are part of the
inflow), they can significantly affect the performance of wetlands designed for the
control of non-point source runoff. A variable hydroperiod, which exhibits dry peri-
ods interspersed with flooding, is a natural cycle in non-point source wetlands, and
fluctuating water levels should be considered a natural feature. A fluctuating water
level could provide needed oxidation of organic sediments and can, in some cases,
rejuvenate a system to higher levels of chemical retention. There was a definitive sea-
sonal cycle, plus sudden bursts of water levels during winter and spring storms at an
agricultural wetland in Ohio described by Fink and Mitsch (2004). Furthermore, not-
ing the importance of biology in wetland types, the water level in one basin in that
wetland dropped almost 30 cm because of burrowing activity by muskrats {Ondatra
zibethicus).

Hydraulic Loading Rate

The hydraulic loading rate, one of the most important variables in treatment wedands,
is defined as:

^=I00Q/A (19. 1)

where

q = hydraulic loading rate (HLR), cm day“^

Q = inflow rate, m^ day“^

A = wetland surface area, m^

Table 19.3 summarizes several recommendations and measurements of HLR of
surface-flow and subsurface-flow wastewater wetlands. Loading rates to surface-flow
wetlands for small municipalities range from 1.4 to 22 cm day“\ while rates to
subsurface-flow constructed wetlands vary between 1.3 and 26 cm day”^. Knight
(1990) reviewed several dozen wetlands constructed for wastewater treatment and
recommended an HLR of 2.5 to 5cm/day for surface-flow constructed wetlands

670 Chapter 19 Wetlands and Water Quality

Table 19.3 Recommended and actual hydrologic loading rates (HLR) for
treatment wetiands

Recommended

Median

Loading Rate,

Type of Wetland

cm/day ^

Surface-flow treatment wetlands

2.5 -5.0

5.4 ± 1.7 (n = 15)

3.0 (n = 205)

Subsurface-flow treatment wetlands

6.0 -8.0

7.5 ± 1.0 (n = 23)

6.8 (n = 634)

and 6 to 8 cm day“^ for subsurface-flow wetlands. Kadlec (2009a) followed up with
a review of more than 800 treatment wetlands 20 years later and found the median
numbers for both surface-flow and subsurface-flow treatment wedands to be almost
exactly midpoint between those ranges.

Detention Time

Detention time of treatment wetlands is calculated as:

t=VplQ^ (19.2)

where

t = theoretical detention time, day

V = volume of wetland basin, m^ (volume of water column for
surface-flow wetlands; volume of medium for subsurface-flow
wetlands

p = porosity of medium (e.g., sand or gravel for subsurface-flow
wedands)

= 1.0 for surface-flow wetlands
Q^= flow rate through wedand, m^ day"'^

The optimum detention dme (or nominal residence time) has been suggested to
be from 5 to 14 days for treatment of municipal wastewater. Florida regulations on
wedands require that the volume in the permanent pools of the wedand must provide
for a residence dme of at least 14 days. Calculation of detendon dme or nominal
residence dme with Equation (19.2) is not always realisdc because of short-circuidng
and the ineffective spreading of the waters as they pass through the wetland. Tracer
studies of flow through wetlands have illustrated the importance of not overrelying on
the theoredcal detendon dme to design treatment wedands. Not all parcels of water
that enter at a certain dme leave the wetland at the same time. In some instances, water
will short-circuit through the wedand, whereas other water will remain in backwater
locations for considerably more dme than the theoretical detention time.

Basin Morphology

Several aspects related to the morphology of constructed wedand basins need to be
considered when designing wetlands. For example, Florida reguladons for the Orlando

Water Quality Wetland Design 671

area require, for littoral zones, a shelf with a gentle slope of 6:1 or flatter to a point
of from 60 to 77 cm below the water surface. Slopes of 10:1 or flatter are even bet-
ter. A flat littoral zone maximizes the area of appropriate water depth for emergent
plants, thus allowing more wetland plants to develop more quickly and allowing wider
bands of different plant communities. Plants will also have room to move “uphill” if
water levels are raised in the basins because of flows being higher than predicted or
to enhance treatment. Bottom slopes of less than 1 percent are recommended for
wetlands built to control runoff, whereas a substrate slope, from inlet to outlet, of
0.5 percent or less has been recommended for surface-flow wetlands used to treat
wastewater.

Flow conditions should be designed so that the entire wetland is effective in
nutrient and sediment retention if these are desired objectives. This may necessitate
several inflow locations and a wedand configuration to avoid channelization of flows.
A length-to-width ratio (L/W) (called the aspect ratio) should be at least 10:1 if water
is purposely introduced to the system. A minimum aspect ratio of 2:1 to 3:1 has been
recommended for surface-flow wastewater wetlands.

Providing a variety of deep and shallow areas is optimum. Deep areas (>50cm),
while too deep for continuous emergent vegetation, offer habitat for fish, increase
the capacity of the wetland to retain sediments, can enhance nitrification as a prelude
to later denitrification if nitrogen removal is desired, and can provide low-velocity
areas where water flow can be redistributed. Shallow depths (<50cm) provide maxi-
mum soil-water contact for certain chemical reactions, such as denitrification, and can
accommodate a greater variety of emergent vascular plants.

Individual wetland cells, placed in series or parallel, often offer an effective design
to create different habitats or establish different functions. Cells can be parallel so that
alternate drawdowns can be accomplished for mosquito control or redox enhance-
ment, or they can be in a series to enhance biological processes.

Chemical Loadings

When water flows into a wetland, it brings chemicals that may be beneficial or possi-
bly detrimental to the functioning of that wetland. In an agricultural watershed, this
inflow will include nutrients such as nitrogen and phosphorus as well as sediments
and possibly pesticides. Wetlands in urban areas can have all of these chemicals plus
other contaminants such as oils and salts. Wastewater, when added to wetlands, has
high concentrations of nutrients and, with incomplete primary treatment, high con-
centrations of organic matter (BOD) and suspended solids. At one time or another,
wetlands have been subjected to all of these chemicals, and they often serve as effective
sinks. Wetlands can be sized using design graphs, standard retention rates, or empirical
models.

Design Graphs

The simplest model available to estimate the retention of nutrients or other chemicals
by wetlands is to use design graphs that give some measure of chemical retention versus
chemical loading, either areal (e.g., g m“^ yr~^) or volumetric (e.g., g m“^ yr~^)- If ^

672 Chapter 19 Wetlands and Water Quality

Figure 19.9 Decrease in nitrate-nitregen by (a) mass, and (b) concentratien for created
and managed wetiands in the Mississippi River Basin. Each data point represents data for a
compiete year for a wetland. Outside lines are 95 percent confidence intervals. Vertical lines
in graphs indicate median loading rate of 60g-N m~^ yr~^. (Mitsch et al., 2005b, copyright
Elsevier, reprinted with permission)

wetland were designed to retain nutrients, for example, it would be desirable to know
how well that retention would occur for various nutrient inflows. Data compiled from
a large number of wetland sites in North America and Europe provide an indication of
the nutrient retention of wetlands. For example. Figure 19.9, compiled from wetlands
in the Mississippi River Basin, illustrates the percentage removal of nitrate-nitrogen
versus loading for the midwestern United States in two ways: (1) mass retention per
unit area, and (2) percentage retention by concentration. Each of the data points is
based on one year’s data at one wetland basin in either the Midwestern United States
or the river delta in Eouisiana.

Water Quality Wetland Design 673

Table 19.4 Nutrient and sediment removal rates and efficiency in constructed wastewater
wetlands

Wetland Type Parameter

Loading
(g m-2 yr-i)

Retention
(g m-2 yr-i)

Percentage

Retention

SURFACE-FLOW CONSTRUCTED WETLANDS

Nitrate -i- nitrate nitrogen

44.4

Total nitrogen

277

126

45.6

Total phosphorus

4.7-56

2.1-45

46-80

Suspended solids

107-6,520

65-5,570

61-98

SUBSURFACE-FLOW CONSTRUCTED WETLANDS

Nitrate -i- nitrate nitrogen

5,767

547

9.4

Total nitrogen

1,058

569

53.8

Total phosphorus

131-631

11-540

8-89

Suspended solids

1,500-5,880

1,100-4,930

49-89

Source’. Kadlec and Knight, 1996.

Retention Rates

Another approach to estimating the retention of nutrients is to simply compare sev-
eral studies and estimate the chemical retention that consistently happens in wet-
lands. Averages from data from many constructed wastewater wedands are shown
in Table 19.4. In general, as suggested by the HLR data in Table 19.3, subsurface
wetlands receive more wastewater and thus receive greater loadings of chemicals and
sediments. The high average mass retention of nitrate-nitrogen in subsurface wet-
lands is due more to these high loading rates in subsurface-flow wetlands than it
is to any ability of these systems to sequester more nitrate-nitrogen. Note that the
percent nitrate-nitrogen retention is much higher in the surface-flow than in the
subsurface-flow wetlands. The retention of phosphorus tends to be more variable in
subsurface wetlands than in surface wetlands.

Summaries of retention rates for several wetlands intercepting non-point source
pollution are given in Table 19.5. Rules of thumb for this type of wetland are that
wetlands can consistently retain phosphorus in amounts of 0.5 to 5 g-P m”^ yr“^and
nitrogen in amounts of about 10 to 40 g-N yr~^ (Mitsch et ah, 2000). To main-

tain biological diversity in the plant community, the lower end of these loading rates
should be used. The long-term rates for phosphorus and nitrogen for the 30-year-old
Houghton Lake treatment wetland described in Case Study 1 in this chapter and
the phosphorus retention rates for the stormwater urban and agricultural wetlands in
south Florida described in Case Studies 2 and 3 fit in these ranges. Nitrate-nitrogen
retention capabilities of freshwater marshes receiving non-point source pollution in
seasonal to cold climates shows a range of nitrogen retention from 3 to 93 g-N m~^
yr“^ and a phosphorus retention rate of 0.1 to 6 g-P yr“^ (Table 19.5) Low
retention numbers are generally from wetlands that are “underfed” nutrients. High
numbers are usually only periodic and therefore would be inappropriate to use for
design purposes.

Table 19.5 Nutrient retention in constructed and natural wetlands receiving
low-concentration (i.e., non-wastewater, nutrient ioading from rivers, overfiows, or non-point
source pollution)

Wetland Location

Wetland

Nitrogen

Phosphorus

Reference

and Type

Size, ha

g-N m"2 yr^

g-P m"2 yr“i

WARM CLIMATE
Everglades marsh, S. Elorida

8000

I0.4-O.6

Richardson and Craft,

1993; Richardson et al.,
1997

Elorida Everglades

23,000

—

1.25

See Case 2, this chapter

stormwater treatment

areas (STAs)

Boney Marsh, S. Elorida

4.9

0.36

Moustafa et al., 1996

Everglades Nutrient Removal

1545

10.8

0.94

Moustafa, 1999

Project, S. Florida
Restored marshes.

3.5

Comin et al., 1997

Mediterranean delta,
Spain

Constructed rural wetland.

0.045

2.8

Raisin et al., 1997

Victoria, Australia
Breton Sound Estuary,

110,000

3.5

Lundberg et al., 2014

Louisiana Delta
COLD CLIMATE
Houghton Lake, Michigan

100

24.39

1.76

Kadlec, 2009b

(30 years)

Constructed wetlands, NE

Phipps and Crumpton,

Illinois

1994; Mitsch et al.,
1995

river-fed and high-flow

211-38

1.4-2. 9

river-fed and low-flow

2-3

23-13

0.4-1. 7

Artificially flooded meadows.

180

43-46

—

Leonardson et al., 1994

southern Sweden

Constructed wetland basins.

0.035-0.09

50-285

26-71

Braskerud, 2002a, b

Norway

Palustrine freshwater

Reinhelt and Horner, 1995

wetlands, NW Washington
urban area

0.44

rural area

—

3.0

Created instream wetland.

—

2.9

Niswander and Mitsch,

1995

Created riverine wetlands.

38.8

2.4

Mitsch et al., 1998, 2014;

Spieles and Mitsch,
2000a; Nairn and
Mitsch, 2000

Created river diversion

4.5

Fink and Mitsch, 2007;

wetland, OH

Mitsch et al., 2008

Agricultural wetlands, OH

1.2

239

6.2

Fink and Mitsch, 2004

Agricultural wetlands, IL (3)

0.3-0.8

0.1

Kovacic et al., 2000

Natural marsh, Alberta,

360

—

iQ.43

White et al., 2000

Canada

^estimated by phosphorus accumulation in soil
^nitrate-nitrogen only

674

Water Quality Wetland Design 675

Empirical Models

A third method for estimating the ability of wetlands to retain chemicals is to use
equations that either have a theoretical base or are empirically determined from large
databases of existing wastewater wedands. One such general model, originally devel-
oped by Kadlec and Knight (1996) and others, is based on a mass-balance approach
called the “k-C* model” and is given as:

qidCldy) = k^(C-C*) (19.3)

where

y = fractional distance from inlet to oudet, unitless
C = chemical concentration, g m“^

= areal removal rate constant, m yr~^

C* = residual or background chemical concentration, g m~^

This equation is based on an assumption that processes can be described on
an areal basis. Thus, the coefficient has units of velocity and can be recognized
as being similar to a settling velocity coefficient used in sedimentation models. C*
represents a background concentration of a chemical or constituent, below which
it is generally agreed that treatment wetlands cannot go. Integrating this equation
over the entire length of the wetland, the solution can be expressed as a first-order
areal model:

[(Co-C*)/(Q-C*)] (19.4)

where

Q, = outflow concentration, g m“^

C; = inflow concentration, g m~^
q = hydraulic loading rate, m yr“^

Estimates of the two parameters needed for this model, C* and are listed
in Table 19.6. This equation does not work equally well for all parameters, but it
does provide a way of estimating the area of a wedand necessary for achieving a cer-
tain removal. Rearranging equations 19.4 and 19.1 gives the following calculation of
wetland area for given results:

^ = Qln[(QrC*)/iCrC*)]/k^ (19.5)

where

Q= flow rate through wetland, m^ yr ^

Where this model is insufficiently backed with good data or does not work prop-
erly, stricdy empirical relationships of the outflow concentration C„ as a function of
the inflow concentration Q and the hydraulic loading rate ( q) have been developed
(Table 19.7).

676 Chapter 19 Wetlands and Water Quality

Table 19.6 Parameters for first-order areal model given in equations to for
several constituents of wastewater wetlands (Subsurface-flow constructed
wetlands and surface-flow constructed wetlands are given as wetland type
where appropriate.)

Constituent and Wetland Type

(m yr

C* (g m 3)

BOD, surface-flow

3.5 -1- 0.053Cj

BOD, subsurface-flow

180

3.5 -1- 0.053Ci

Suspended solids, surface-flow

1,000

5.1 -1- o.ieq

Total phosphorus, surface and subsurface-flow

0.02

Total nitrogen, surface-flow

1.5

Total nitrogen, subsurface-flow

1.5

Ammonia nitrogen, surface-flow

Ammonia nitrogen, subsurface-flow

Nitrate nitrogen, surface-flow

Nitrate nitrogen, subsurface-flow

Source: Kadlec and Knight, 1996.

Other Chemicals

Although most evaluations of the efficiency of wetlands have been concerned with this
capacity to remove nutrients, sediments, and organic carbon (BOD), there is some
literature on other chemicals, such as iron, cadmium, manganese, chromium, copper,
lead, mercury, nickel, and zinc. Wetland soils or biota or both often easily sequester
metals. That is the basic problem in using wetlands as sinks tor such chemicals: They
can accumulate in the food chain.

Soils

The topsoil is important to the overall function of a constructed wetland (Fig. 19.10).
It is the primary medium supporting rooted vegetation, and, particularly for subsur-
face wedands, it is part of the treatment system. The sediments retain certain chemicals
and provide the habitat for micro- and macroflora and fauna that are involved in chem-
ical transformations. Constructed wetland soil texture depends on whether surface
flow over the substrate or subsurface flow through the substrate is being considered.
Surface-flow wetland soils are generally less effective in removing pollutants per unit
area but are closer in design to natural wetlands. Their ability to provide structure and
nutrition to the wetland plants is important. Clay material, although favored as a sub-
surface liner, limits root and rhizome penetration and may prevent water from reaching
plant roots. Silt clay or loam soils are preferable for the overlying soils in constructed
wetlands. Sandy soil is less preferred for surface -flow wetlands. For subsurface-flow
wetlands, high permeability is preferred. The material needs to be sand, gravel, or
some other highly permeable media.

The subsoil of constructed wedands (usually below the root zone and referred to
as a liner) must have permeability low enough to cause standing water or saturated

Water Quality Wetland Design 677

Table 19.7 Empirical equations for the estimation of outfiow concentrations or
wetiand area based on inflow concentrations and hydrauiic retention time
(Correlation coefficient (R^) and number of wetiands used in anaiysis (n) are aiso
given.). Cj, inflow concentration (g m~^); C„, outfiow concentration (g m~^); A, area
of wetiand (ha); Q, wetiand inflow, (m^/day); q, hydraulic ioading rate, (cm/day).

Constituent

Equation®

r2 (n)

BOD

Surface-flow wetlands

= 4.7 -1- 0.17Sq

0.62

(440)

Subsurface-flow, soil

= 1.87 -1- O.liq

0.74

(73)

Subsurface-flow, gravel

= 1.4 -1- o.ssq

0.48

(100)

Suspended solids

Surface-flow wetlands

= 5.1 -1- 0.158C,

0.23

(1,582)

Subsurface-flow wetlands

= 4.7 -1- 0.09C|

0.67

(77)

Ammenia nitrogen

Surface-flow wetlands
Surface-flow marshes

A =
Co

= 0.01Q/exp[1.527 In C„ -1.05 In q -I- 1.69]

= 0.336C°''^28q0.456

0.44

(542)

Subsurface-flow wetlands

= 3.3 -1- 0.46C,

0.63

(92)

Nitrate nitrogen

Surface-flow marshes

0.35

(553)

Subsurface-flow wetlands

= 0.62q

0.80

(95)

Total nitrogen

Surface-flow marshes

= 0.409q -1- 0.122^

0.48

(408)

Subsurface-flow wetlands

= 2.6 -1- 0.46C, -1- 0.124q

0.45

(135)

Total phosphorus

Surface-flow marshes

= 0.195C°-®^q°'®®

0.77

(373)

Surface-flow swamps

= 0.37C°''^°tj°®®

0.33

(166)

Surface-flow wetlands

= 0.51C>-“

0.64

(90)

soils. If clay is not available on site, it may be advisable to add a layer of clay to min-
imize percolation. Studies have also been undertaken to investigate other materials
as liners for constructed wetlands. The most frequently used liners for constructed
wetlands are clays, clay bentonite mixtures, or synthetic materials, such as polyvinyl
chloride (PVC) and high-density polyethylene (HDPE). Experiments have been con-
ducted in recycling materials such as coal combustion waste products. As it turns out,
using calcium-rich sulfur-scrubber waste material was shown to actually increase the
phosphorus-retention capability of the wetlands (Ahn et ah, 2001; Ahn and Mitsch,
2001), but care must be taken that the material completely seals the wetland because
leachate from this liner material is highly alkaline.

Subsurface flow through subsurface wetlands can be through soil media {root-zone
method) or through rocks, gravel, or sand {rock-reed filters). Flow in both cases is 15 to
30 cm below the surface. Gravel is sometimes added to the substrate of subsurface-flow
wetlands {gmvel-hed) to provide a relatively high permeability that allows water to
percolate into the root zone of the plants where microbial activity is high. Gravel

678 Chapter 19 Wetlands and Water Quality

can be silica based or limestone based; the former has less capacity for phosphorus
retention. Another evaluation of the European-design subsurface wetlands indicated
that they often decrease in hydrologic conductivity after several years and become
clogged, essentially becoming partial-surface-flow wetlands.

Organic Content

The organic content of soils has some significance for the retention of chemicals in
a wetland. Mineral soils generally have lower cation exchange capacity than organic
soils do; the former is dominated by various metal cations, and the latter is dominated
by the hydrogen ion. Organic soils can therefore remove some contaminants (e.g.,
certain metals) through ion exchange and can enhance nitrogen removal by providing
an energy source and anaerobic conditions appropriate for denitrification. Organic
matter in wetland soils varies between 5 and 75 percent, with higher concentrations in
peat-building systems, such as bogs and fens, and lower concentrations in mineral-soil
wetlands, such as riparian bottomland wetlands subject to mineral sedimentation or
erosion. When wetlands are constructed, especially subsurface-flow wetlands, organic
matter, such as composted mushrooms, peat, or detritus, is often added in one of the
layers. For construction of many wetlands, however, organic soils are avoided because
they are low in nutrients, can cause low pH, and often provide inadequate support for
rooted aquatic plants.

Depth and Layering of Soil

The depth of substrate is an important design consideration for wastewater wetlands,
particularly those that use subsurface flow. The depth of suitable topsoil or substrate
should be adequate to support and hold vegetation roots. A common substrate depth
for constructed wetlands is 60 to 100 cm. In some cases, layering more elaborate than
that shown in Figure 19.10 is suggested.

Soil Chemistry

Although exact specifications of nutrient conditions in a wedand soil necessary to
support aquatic plants are not well known, low nutrient levels characteristic of organic,
clay, or sandy soils can cause problems for initial plant growth. Although fertilization
may be necessary in some cases to establish plants and enhance growth, it should
be avoided if possible in wetlands that eventually will be used as sinks for the same
macronutrients. When fertilization is required to get plants started in constructed
wetlands, slow-release granular and tablet fertilizers are often useful.

When soils are submerged and anoxic conditions result, iron is reduced from ferric
(Pe+++) to the ferrous (Fe'*’'*’) ions, releasing phosphorus that was previously held as
insoluble ferric phosphate compounds. The Fe-P compound can be a significant source
of phosphorus to overlying and interstitial waters after flooding and anaerobic condi-
tions occur, particularly if the wetland was constructed on previously agricultural land.
After an initial pulse of released phosphorus in such constructed wetlands, the iron
and aluminum contents of a wetland soil exert significant influences on the ability of
that wetland to retain phosphorus. All things being equal, soils with higher aluminum

Water Quality Wetland Design 679

Figure 19.10 Soil cross-sections of (a) surface-flow wetland, and (b) subsurface-finw
wetland. (After Knight, 1990)

and iron concentrations are more desirable because their affinity for phosphorus
is higher.

Vegetation

Just as the question “What plants should be used?” arises for creating and restor-
ing wetlands as discussed in Chapter 18, vegetation choice is also a consideration
for treatment wetlands. But there is at least one significant difference for treatment
wetlands: While creation and restoration of wetlands are done principally to develop
a diverse vegetation cover and provide habitat, treatment wetlands are constructed
with the main goal of improving water quality. The plants in created and restored
wetlands are part of the solution; in treatment wetlands, they are the partial cause of

680 Chapter 19 Wetlands and Water Quality

the solution. Furthermore, treatment wetlands invariably have higher concentrations
of chemicals in the water, which by its very nature limits the number of plant species
that will survive in those wetlands. Experience has shown that relatively few plants
thrive in the high-nutrient, high-BOD wastewaters that are applied to treatment wet-
lands. Vymazal (2013) found that a total of 150 different species of plants were used in
643 surface-water wetlands from 43 countries. Table 18.4 flagged some of the macro-
phyte species that are frequendy used for treatment wetlands amid the hundreds that
are used for creation of wedand habitat. Among those plants are cattails ( Typha spp. ),
the bulrushes {Schoenoplectusspp., Scirpusspp.), and reed grass {Phra-^mites australis).
The last is the preferred plant in subsurface-flow wedands around the world but is not
favored in many parts of North America because of its aggressive behavior in fresh-
water and brackish marshes. Other commonly used plants included Juncus effusus,
Ekocharis spp., Phalaris arundinacea., and Cyperus papyrus, the last mosdy in Africa.

When water is deeper than 30 cm, emergent plants often have difficulty growing.
In these cases, surface-flow wedands can become covered with duckweed {Lemna
spp.) in temperate zones and water hyacinths [Eichhornia crassipes) and water lettuce
{Pistia spp.) in the subtropics and tropics. While rooted floadng aquadcs, such as
Nymphaea, Nuphar, and Nelumbo, are favored for their aesthedcs, they thrive only in
rare instances in treatment wetlands, where, due to high-nutrient condidons, they are
easily overwhelmed by duckweed and filamentous algae.

Tanner (1996) compared reladve nutrient uptake and pollutant removal of
eight macrophytes in gravel-bed wetland mesocosms fed by dairy wastes in New
Zealand (Fig. 19.11). Greatest above-ground biomass was seen in this highly polluted
wastewater by Glyceria maxima and Zizania latifolia, while greatest below-ground
biomass was seen with Bolboschoenus fluviatilis ( Scirpus fluviatilis in the United
States), which had below- ground biomass 3.3 dmes its above-ground biomass
(Fig. 19.11a). Total nitrogen removed from these mesocosms was linearly correlated
with total plant biomass (Fig. 19.11b). Based on key growth characterisdcs of the
plants in this wastewater, three producdve gramminoids {Zizania latifolia, Glyceria
maxima, and Phraptmites australis) had the highest overall scores. Baumea articulata,
Cyperus involucratus, and Schoenoplectus validus had medium scores, while Scirpus
fluviatilis and Juncus effusus had the lowest scores and are least likely to be effective
plants in wastewater wedands.

Establishing Vegetation

Vegetation can be established through the same general procedures oudined in
Chapter 18 — that is, by planting roots and rhizomes directly or by seeding. Because
these wetlands are usually constructed on former upland with no connection to rivers
or streams, reliance on nature bringing in plant propagules generally does not work.
Field-harvested plants or nursery-grown stock can be used for plantings. The former
have the advantage of establishing vegetation cover more quickly than would smaller
nursery stock. Also, these plants, if harvested nearby, are adapted to the local climate
and may be from the proper genotype for the region. Conversely, harvesting plants
in large numbers from natural wetlands may threaten those wetlands. Nursery plants

4,000

2,000

4,000 - I belowground biomasT

% .A.

Figure 19.11 Results of a study comparing eight macrophytes commonly used in wastewa-
ter wetlands in New Zealand after 124 days of culture in dairy farm wastewater, (a) Mean
above-ground and below-ground biomass accumulation of the eight macrophytes. Differ-
ent letters indicate significant differences, (b) Total nitrogen removal from ammonium-rich
dairy farm wastewater versus total plant biomass. Regression coefficient -- 0.66. (After
Tanner, 1996)

681

682 Chapter 19 Wetlands and Water Quality

are easier to plant because of their generally small size, and a greater diversity and
number of plants can be obtained from good nurseries. However, it is often unclear
what genetic stock was used to start these plants, and they may not be from stock
adapted to the local climate.

Water, either too much or too little, is the major reason why macrophytes do
not become well established in wetlands constructed for wastewater treatment. When
plants are first establishing themselves, the optimum conditions are moist soils or very
shallow (<5cm) water depth. If water is too deep, the new plants will be flooded
out. If there is inadequate water and topsoil dries out, the plants will not survive.
If the wastewater can be used in measured amounts to irrigate the plants, this is
optimum. If not, artificial irrigation might be required to make sure that plants are
successful.

Wetland Management after Construction

Wildlife Control

Although the development of wildlife is a welcomed and often desired aspect of
treatment wetlands, managing plant and animal populations often becomes necessary
maintenance of constructed wedands. In North America, beavers {Castor canadensis)
and muskrats {Ondatra zibethicus) create obstructions to inflows and outflows,
destroy vegetation, or burrow into dikes. (This is one reason why dikes should not
be built up around constructed wetlands if they can be avoided.) Major vegetation
removal, particularly by herbivorous muskrats that use the plant material both for
food and shelter, can turn a fully vegetated marsh into a plant-devoid pond in the
matter of weeks or months. These events are referred to as eat-outs. There is very
little that can be done to prevent these eat-outs except to trap and move the animals,
which is a laborious task.

In other cases, animals such as beaver and muskrat and large birds such as Canada
geese {Branta canadensis) and snow geese {Chen spp.) grazing on newly planted
perennial herbs and seedlings are particularly destructive. The riming of planting is
important, especially if migratory animals are involved in destructive grazing in the
winter. Using gunshot devices and the extract of grape juice as a “hot foot” material
on the adjacent landscape have all been suggested but without any permanent success.
Probably the easiest approach we have noted in many years of observing geese is to
have a wide band of emergent vegetation between where they land (on the water) and
where they like to graze (upland lawns). But, of course, you will have to get the local
muskrats to cooperate and not remove the vegetation.

Similarly, deeper wetlands often become havens for undesirable fish, such as carp
{Cyprinus carpio), which can cause excessive turbidity and uproot vegetation. Carni-
vores such as northern pike {Esox Indus) have been discussed as a potential control of
carp. Total removal of fish by drawdown is probably necessary if carp begin to degrade
outflow water quality excessively. The problem is that this fish removal might affect
mosquito control (see the following discussion).

Wetland Management after Construction 683

Attracting Wildlife

Just as many animals can cause maintenance headaches, the attraction of wildlife to
constructed wetlands is one of the reasons why public support for such projects can be
high in the first place. So every attempt should be made to have a diverse ecosystem
and not just a pond with water flowing through it. Weller (1994) recommends a 50:50
ratio of open water to vegetation cover in marshes to attract water birds, and this ratio,
with proper development of the initial bathymetry of the ponds, is quite easy. Also,
creating diverse habitats with live and dead vegetation, islands, and floating structures
is desirable.

In many cases of wetland construction, wildlife enhancement begins soon after
construction. At a constructed wetland at Pintail Lake in Arizona, the area’s water-
fowl population increased dramatically by the second year of use; duck nest density
increased 97 percent over the first year (Wilhelm et al., 1989). A considerable increase
in avian activity was also noted at the Des Plaines River Wetlands Demonstration
Project in northeastern Illinois. Migrating waterfowl increased from 3 to 15 species
and from 13 to 617 individuals between 1985 (preconstruction) and 1990 (one year
after water was introduced to the wetlands) . The number of wetland-dependent breed-
ing birds increased from 8 to 17 species, and two state-designated endangered birds,
the least bittern and the yellow-headed blackbird, nested at the site after wetland con-
struction (Hickman and Mosca, 1991). At the Olentangy River Wetland Research
Park in Ohio, a total of 174 bird species were listed for the site approximately 15 years
after the initial wetlands were created. This count was high because part of the site
included approximately 7 ha of bottomland forest on its eastern edge and an urban
tree -dominated cemetery was on its western border.

One interesting question related to attracting birds to treatment wetlands is
whether the birds might have an effect on the treatment capacity of the wetlands
and, specifically, if birds are in high numbers, whether their excreta could undermine
the effectiveness of the wetlands for nutrient and organic removals. Anderson et al.
(2003) presented a several-year study done on a 10-ha treatment wedand in northern
California on that possible effect. Bird use peaked at 12,000 individuals during the
second year during a four-month period. Average daily inputs by birds of 2.5 g N
m“^ and 0.9 g P m~^ were found, which represented less than 10 percent of the mean
daily loading rates to the wedand. They concluded that bird use “does not lead to a
significant reducdon in treatment performance.”

Mosquito Control

The subject of mosquito control will always be brought up when wetlands are being
constructed, pardcularly when the wetlands receive runoff or wastewater. In general,
it has been concluded that properly managed wastewater treatment wedands pose no
more mosquito threat than do natural wedands (Knight et al., 2003). Mosquitoes can
be controlled in constructed wedands by changing the hydrologic condidons of the
wedands to inhibit mosquito larvae development (flowthrough condidons discour-
age mosquitoes) or by using chemical or biological control. Many researchers have

684 Chapter 19 Wetlands and Water Quality

proposed mosquito control by fish, especially the air-gulping mosquito fish {Gambu-
sia affinis) or similar small fish. One reason to maintain some deeper areas in wedands
in temperate zones is to allow fish such as Gambusia and other top minnows and
sunfish to survive the winter and feed on mosquito larvae. Little is known about
the role that water quality has on encouraging or discouraging mosquitoes directly,
but the effect of poor water quality by removing fish can have a dramatic effect in
causing mosquito population increases. Bacterial insecticides (e.g.. Bacillus sphaeri-
cus and Bacillus thurinpiiensis var. ismdensis) and the fimgus Lap/enidium gipfanteum
are known pathogens of mosquito larvae, but they have not been tested extensively,
and there is always the possibility of resistance induction in mosquitoes (Knight et al.
2003). Constructing boxes to encourage nesting by swallows (Hirundinidae), swifts
(Apodidae), and bats (Chiroptera) have also been used to control adult mosquito
populations at constructed wetlands.

Some studies have evaluated the relative importance of different macrophyte
species on the propensity of mosquito survival. In general, the denser the plant stands,
the more difficult it is for both predators and mosquito control efforts to reach
the mosquitoes. Thus, the highly productive plants {Typha, several Schoenoplectus
= Scirpus, Phragmites^ and Eichhornia crassipes) have the highest mosquito scores
in Table 19.8. Knight et al. (2003) suggested the following strategy for treatment
wetland design to minimize mosquitoes:

■ Select plant species that optimize both wastewater treatment performance and
mosquito control.

■ Include deepwater zones that are free of emergent and aquatic plants to
provide fish habitat and access to vegetated areas.

■ Limit the width of emergent plant zones to facilitate access by predaceous fish
and for application of chemical control agents.

■ Design wetlands with steep embankments. (Although this is effective for
mosquito control, it is not a good strategy to develop a diverse littoral zone
around the wetland.)

Pathogens

Because many treatment wedands are built specifically to deal with human and ani-
mal wastewater, proper sanitary engineering techniques should be used to minimize
human exposure to pathogens. Treatment wedands are meant to be biologically rich
systems, and microbial acdvity is a major part of the treatment process. Measure-
ments of indicator organisms, such as fecal and total coliforms, should be part of the
monitoring of municipal wastewater treatment wedands. Nearby wells should also be
sampled, because water seeping from a wastewater wedand near potable water sup-
plies should be monitored carefully. If a wetland is being used as terdary treatment
to a convendonal treatment plant, the design of the disinfection system in the con-
ventional treatment plant must be considered. Chlorine disinfection and the resuldng

Table 19.8 Estimated mosquito production propensity of various wetiand piant species.
Scores less than 9 indicate minimal mosquito breeding problems; scores between 9 and 13
indicate a need to maintain a low coverage for this piant species; and scores of 14 and above
indicate a need to minimize the occurrence of the plant species in the wetiand to avoid
mosquito issues.

Plant Group Plant Species

Common Name

Mosquito Production Score^

Rooted emergent piants

Atisma geyeri

Water-plantain

Alisma trivale

Water-plantain

Alopercurus howellii

Foxtail

Carex obnupta

Sedge

Carex rostrata

Sedge

Carex stipata

Sedge

Cyperus aristatus

Flat sedge

Cyperus difformis

Flat sedge

Cyperus esculentus

Flat sedge

Cyperus niger

Flat sedge

Deschampsia danthonides

Grass

Echinochloa crusgalti

Barnyard grass

Echinodorus berteroi

Burhead

Eleocharis palustris

Spikerush

Equisetum arvense

Florsetall

Erankenia grandifolia

Alkali heath

Glyceria leptostachya

Mannagrass

Juncus acutus

Softrush

Juncus effusus

Softrush

Jussiaea repens

Primrose

Leers/a oryzoides

Rice cutgrass

Leptochloa fasicularis

Salt-meadow grass

Ludwigia spp.

Primrose willow

Lythrum californicum

Loosestrife

Oryza sativa

Rice

Phalaris arundinacea

Reed canary grass

Phragmites australis

Common reed

Plantago major

Common plantain

Polygonum amphibium

Water smartweed

Polygonum hydropiperoides

Smartweed

Polygonum pennsylvanicum

PInkweed

Polygonum punctatum

Smartweed

Polypogon elongatus

Rabbitfoot grass

Potentilla palustris

Cinquefoil

nerididum aquilinum

Fern

Sagittaria latifolia

Duck-potato

Sagittaria longiloba

Arrowhead

Sagittaria montevidensis

Giant arrowhead

Scirpus acutus

Bulrush

Scirpus americanus

Three-square bulrush

{continued)

685

Table 19.8 (Continued)

Plant Group Plant Species

Common Name

Mosquito Production Score^

Scirpus californicus

Giant bulrush

Scirpus oineyi

Alkali bulrush

Sparganium eurycarpum

Burreed

Typha angustifolia

Narrowleaf cattail

Typha glauca

Cattail

Typha latifolia

Common cattail

Zizania aquatica

Wildrice

Floating aquatic plants

Azolia fiiicuioides

Water fern

Bacopa nobsiana

Water hyssop

Brasenia schreberi

Water shield

Eichhornia crassipes

Water hyacinth

Hydrocotyie ranuncuioides

Pennywort

Hydrocotyie umbeilata

Pennywort

Lemna gibba

Duckweed

Lemna minima

Duckweed

Nasturtium officinale

Water cress

Nuphar poiysepaium

Spatterdock

Pistia stratiotes

Water lettuce

Potamogeton crispus

Curled pondweed

Potamogeton diversifolius

Pondweed

Ranunculus aquatilis

Buttercup

Ranunculus flammula

Buttercup

Spirodela polyhylza

Duckmeat

Wolffiella lingulata

Bog mat

Submerged aquatic piants

Callltriche longipedunculata

Water starwort

Ceratophyllum demersum

Coontail

Eleocharis acicularis

Spikerush

Elodea canadensis

Waterweed

Elodea densa

Waterweed

Isoetes howellli

Quillwort

Isoetes orcuttii

Quillwort

Lilaeopsis occidentalis

Lilaeosis

Myriophyllum spicatum

Water milfoil

Najas flexilis

Naiad

Najas graminea

Naiad

Potamogeton fiilformis

Pondweed

Potamageton pectinatus

Sago pondweed

Ruppia spiralis

Ditchgrass

Utricularia gibba

Bladderwort

Utricularia vulgaris

Bladderwort

Zannichellia palustris

Horned pondweed

Source’. Knight et al., 2003, and Collins and Resh, 1989.

686

Wetland Management after Construction 687

chlorine residual would cause significant problems in treatment wedands, so other
means of disinfection (ozonation or ultraviolet radiation) should be used if disinfection
is required before the wastewater enters the wetland.

Water-Level Management

The water level of surface-flow treatment wedands is the key to both water quality
enhancement and vegetadon success. Most constructed municipal wastewater wet-
lands have litde control on the overall inflow of wastewater. Flow and depth are first
controlled by designing the basin large enough to create the proper HLR. Most con-
structed wedands have a control structure, such as a flume or weir, to control outflow;
these structures should be flexibly designed so they can be manipulated to control
water depth. Too much water stresses macrophytes as much as too litde water. Water
depths of 30 cm or less are opdmum for most herbaceous macrophytes used in treat-
ment wedands. Water depths greater than 30 cm can lead to vegetation reducdon.

Compounding the effect that water level has on vegetadon is the effect that it has
on wastewater treatment. Deep water favors a high HLR and sediment and phospho-
rus retendon associated with sedimentadon and similar processes; it also leads to less
resuspension, longer retendon dme, more organic matter accumuladon, and lower
redox condidons in the bottom waters. Shallow water leads to closer proximity of
sediments and overlying water and more oxygen in the soil. Optimizing wastewater
treatment and vegetadon success is a continual balancing act.

Greenhouse Gas Emissions

Some concern has been expressed about treatment wedands because of their emission
of greenhouse gases, pardcularly methane (CH4) In a survey of 158 published papers
on constructed wedands, median carbon dioxide (CO2) emissions were significandy
lower in surface -flow treatment wedands (840 g-C m“^ yr^^) than in subsurface flow
wedands (1200 g-C yr^^) (Mander et ah, 2014). Median methane emissions
ranged from 35 g-C rcr^ yr“^ from surface-water wedands to 56 g-C irr^ yr“^ from
horizontal subsurface wetlands. The CO2 emissions from the treatment wetland
soils are probably more than compensated for by vegetation producdvity and soil
carbon sequestration. The methane emissions are comparable to the rates presented
in Chapter 17 (see Table 17.3), where an average methane emission rate of sev-
eral temperate zone wedands was reported to be 58 g-C yr^h Mander et al.
(2014) pointed out that the methane emissions from treatment wedands are one to
two orders of magnitude lower than those coming from convendonal wastewater
treatment plants. Nitrous oxide (N2O) emissions in this study had mediums of 0.8,
1.0, and 1.1 g-N yr^^, respecdvely, for surface-water, verdcal subsurface, and
horizontal subsurface treatment wedands. These low N2O emissions were theorized
by Mander et al. (2014) to be mostly residuals from denitrification, which mosdy
resulted in emissions of inert N2 gas.

688 Chapter 19 Wetlands and Water Quality

Economics and Values of Treatment Wetlands

It is generally believed that treatment wetlands are less expensive to build and maintain
than conventional wastewater treatment, and that is the appeal of these systems to
many people. However, cost comparisons should be made carefully before investing
in these systems. Any estimate of the cost of a new wetland’s development should
include these four items:

1 . Engineering plan

2. Preconstruction site preparation

3. Construction costs (e.g., labor, equipment, materials, supervision, indirect
and overhead charges)

4. Cost of land

Capital Costs

An equation estimating the cost of constructing wedands in general, including several
wedands that are not wastewater wedands (not including the cost of land) is:

where

= 196^='’-®“

(19.6)

Cyi = capital cost of wedand construcdon per unit area,

U.S.$ X 1,000 ha-i
A = area of wetland, ha

This reladonship suggests that a 1-ha wedand would cost almost $200,000, a
10-ha wedand would cost $60,000 per ha, and a 100-ha wedand would cost $19,000
per ha. The data clearly suggest that there is an economy of scale involved in wetland
construction. This equation included all types of created and constructed wedands,
not only treatment wetlands.

Kadlec (2009a) compared the costs of 92 surface water wetlands and 63 horizontal
subsurface wedands and came up with the reladonships shown in Figure 19.12 and
listed here:

surface water wedands C = 194^*’-^^*’ 0.03 ha < < 10, 000 ha (19.7)

and

subsurface wetlands C = 652H®'^®^ 0.005 ha < A < 20 ha (19.8)

where

C = capital cost, U.$.$ X 1,000 (2006 $)

A = area of wetland, ha

Operating and Maintenance Costs

Operating and maintenance costs vary according to the wedand’s use and to the
amount and complexity of mechanical parts and plumbing that the wetland contains.

Economics and Values of Treatment Wetlands 689

Figure 19.12 Costs of treatment wetlands versus wetland size fer surface water wetiands
and herizontai subsurface wetiands based on data from 92 surface water wetlands and 63
subsurface wetlands. (After Kadlec, 2009a)

Fewer data on operational costs are available. Kadlec and Knight (1996) estimated the
operation and maintenance costs for one wastewater wetland to be about $85,500
per year. That estimate included $50,000 per year for personnel to be in charge of
the 175-ha wetland. A wide range of $5,000 to $50,000 per year of operating and
maintenance costs was estimated by those authors from smaller wetlands. Gravity-fed
wetlands are far less expensive to maintain than highly mechanized wedands that need
significant plumbing and pumps. With current monitoring and legal reporting require-
ments for treatment wedands, it is likely that current operating and maintenance costs
would be at least $50,000 to $100,000 per year. This esdmate would probably not
include animal and vector control and unclogging (for subsurface wedands) acdvides,
which can be cosdy (Kadlec, 2009a).

Other Benefits of IVeatment Wetlands

fiubsurface treatment wetlands provide little additional benefit beyond the water
quality improvement they were designed to provide, but surface-flow treatment
wetlands have a variety of additional benefits. The watery habitat that is created can
be a major ancillary benefit of these systems. In addition to providing habitat for
mammals such as nutria, beavers, muskrats, amphibians, fish, and voles surface-flow
treatment wedands are often a haven for waterfowl and wading birds. Human uses,
such as trapping and hundng, are not incompadble with some wastewater wedands.
If designed properly in an urban area, wedands are locadons where the public can
visit and learn about their important water quality role. This message is a powerful
one to the uninitiated, and they often become ardent wetland conservationists as a
result of seeing wetlands at work.

690 Chapter 19 Wetlands and Water Quality

Another benefit of using both natural and constructed wetlands for water quality
improvement relates to areas where land-building is needed. In the subsiding envi-
ronment of Louisiana’s Gulf Coast, nutrients are permanently retained in the peat of
wetlands receiving high-nutrient wastewater as the wetland aggrades to match subsi-
dence. In this case , wastewater discharge into a wetland can occur without saturating the
system and simultaneously helps counteract the deleterious effects of land subsidence.

Comparing Wetlands and Conventional Technology

A comparison of the construction and operating costs of a proposed large (>2,000
ha) wetland that was to be constructed in the Florida Everglades with conventional
chemical treatment is illustrated in Table 19.9. In this example, if land costs are not
considered, the wedand alternative has an II percent lower capital cost and a 56 per-
cent lower operating cost than the chemical treatment alternative. Although land costs
can be significant for treatment wetlands, particularly in urban areas (in essence, solar
energy is being substituted for fossil fuel energy), it is generally not appropriate to use
the cost of land in comparison with technological solutions that require little land.
This is because the land being used by the wetland can be sold after the life of the
wetland is completed, while the salvage value of the worn-out equipment used for
conventional treatment alternatives is generally zero.

One of the more clever calculations of the difference between using wetlands ver-
sus conventional mechanical systems for wastewater treatment is an illustration of the

Table 19.9 Estimated cest comparison for phosphorus control in 760,000 m^ day ^
agricultural runoff wetland in Florida

Treatment Wetland

Chemical Treatment

Land cost

$34,434,000

$2,140,000

Capital costs (land free)

$95,836,000

$108,260,000

Total annual operation/maintenance

$1,094,000

$2,490,000

O&M, present worth

$33,443,000

$76,153,000

Total present worth, without land cost

$129,279,000

$185,637,000

Total present worth, with land cost

$163,713,000

$187,777,000

Source: Kadlec and Knight 1996.

Table 19.10 Net atmospheric generation of carbon for a 3,800 m^ day~^
wastewater treatment facility using treatment wetlands or conventional mechanical
treatment

Treatment Wetland

Conventional Treatment

Carbon Flow

(metric tons C day“^)

Atmospheric carbon from power generation

1,350

Carbon sequestration

-3

Net atmospheric carbon

1,350

Source: Ogden 1999.

Summary Considerations 691

relative impact on the emission of the greenhouse gas CO2 . Normalizing estimates for
a 3,800 m^ day”^ (1 million gal/day) flow of wastewater, mechanical treatment leads
to 27 times more emission of C02to the atmosphere than does a treatment wetland
(Table 19.10). The wetland system, in fact, has the additional benefit of sequestering a
small amount of carbon. Conventional wastewater treatment uses 3.9 kg of fossil fuel
carbon to remove 1 kg of carbon; a wetland treatment system uses 0.16 kg of fossil
fuel carbon to remove 1 kg of carbon (Ogden, 1999).

Summary Considerations

Wastewater treatment wetlands are not the solution to all water quality problems and
should not be viewed as such. Many pollution problems, such as excessive BOD or
metal contamination, may require more conventional approaches. Yet the fact that
thousands of wetlands have been constructed around the world for pollution con-
trol attests to their importance and value. Several considerations, both technical and
institutional, must be considered as treatment wetlands are designed and built.

Technical Considerations

1 . Values of the wetlands, such as wildlife habitat, should be considered in any
treatment wetland development.

2. Acceptable pollutant and hydrologic loadings must be determined for the use
of wetlands in wastewater management. Appropriate loadings, in turn,
determine the size of the wetland to be constructed. Overloading a
constructed wetland can be worse than not building it at all.

3. All existing characteristics of local natural wetlands, including vegetation,
geomorphology, hydrology, and water quality, should be well understood so
that natural wetlands can be “copied” in the construction of treatment
wetlands.

4. Particular care should be taken in the wetland design to address public
health, including mosquito control and protection of groundwater resources.

Institutional Considerations

1 . Wastewater treatment by wetlands often can serve the dual purposes of both
wetland habitat development and wastewater treatment and recycling. The
creation of treatment wetlands as mitigation for lost wetlands is still generally
unacceptable because of the lack of sustainability and the high level of
pollution of treatment wedands compared to restored wetlands, but
opportunities for dual use of wetlands should continue to be explored.

2. Many permit processes in governments do not recognize treatment wetland
systems as alternatives for wastewater treatment. In these cases, experimental
systems should first be established for a given region. Modification of
requirements for granting permits for pilot wetlands is needed to make
effective progress in developing approaches.

692 Chapter 19 Wetlands and Water Quality

It is useful to remember that wetland design is an inexact science and that per-
turbation and biological change are the only things we can be sure of in these created
ecosystems. Traditional engineering approaches to wastewater wetlands, without an
appreciation of self-design in ecosystems, are sure to cause disappointment. If a treat-
ment wetland continues to function according to its main goal — that is, improving
water quality — changes in plant species should not be viewed as that significant unless
invasive nonnative plants become dominant.

Recommended Readings

IWA Specialists Group on Use of Macrophytes in Water Pollution Control. 2000.
Constructed Wetlands for Pollution Control. Scientific and Technical Report No.
8. London: International Water Association (IWA).

Kadlec, R., and S. Wallace. 2009. Treatment Wetlands., 2nd ed. Boca Raton, FL: CRC
Press.

References

Ahn, C., and W. J. Mitsch. 2001. Chemical analysis of soil and leachate from experi-
mental wetland mesocosms lined with coal combustion products. Journal of Envi-
ronmental Quality 30: 1457-1463.

Ahn, C., W. J. Mitsch, and W. E. Wolfe. 2001. Effects of recycled FGD liner material
on water quality and macrophytes of constructed wetlands: A mesocosm experi-
ment. Water Research iS'. 633-642.

Anderson, D. C., J. J. Sartoris, J. S. Thullen, and P. G. Reusch. 2003. The effects
of bird use on nutrient removal in a constructed wastewater-treatment wetland.
Wetlands 23: 423M:35.

Arheimer, B., and H. B. Wittgren. 1994. Modelling the effects of wetlands on regional
nitrogen transport. Amhio2Z\ 378-386.

Boustany, R. G., C. R. Crozier, J. M. Rybczyk, and R. R. Twilley. 1997. Denitri-
fication in a south Louisiana wetland forest receiving treated sewage effluent.
Wetlands Ecology and Management 4: 273-283.

Braskerud, B. C. 2002a. Factors affecting nitrogen retention in small constructed
wetiands treating agricultural non-point source pollution. Ecological Engineering
18: 351-370.

Braskerud, B. C. 2002b. Factors affecting phosphorus retention in small constructed
wetlands treating agricultural non-point source pollution. Ecological Engineering
19: 41-61.

Brix, H. 1987. Treatment of wastewater in the rhizosphere of wedand plants: The
root-zone method. Water Science and Technology 19: I07-II8.

Brix, H. 1994. Use of constructed wedands in water polludon control: Historical
development, present status, and future perspectives. Water Science and Technology
30: 209-223.

Brix, H. 1998. Denmark. In J. Vymazal, H. Brix, P. F. Gooper, M. D. Green, and
R. Haberl, eds. Constructed Wetlands for Wastewater Treatment in Europe. Back-
huys Publishers, Leiden, the Netherlands, pp. I23-I52.

References 693

Brix, H., and H.-H. Schierup. 1989a. The use of aquatic macrophytes in water- pol-
lution control. Ambio 18: 100-107.

Brix, H., and H.-H. Schierup. 1989b. Sewage treatment in constructed reed
beds — Danish experiences. Water Science and Technology 21 ■. 1655-1668

Brix, H., and C. A. Arias. 2005. The use of vertical flow constructed wetlands for
on-site treatment of domestic wastewater: New Danish guidelines. Ecolo£iical
Engineering 25: 491-500.

Brodie, G. A., D. A. Hammer, and D. A. Tomljanovich. 1988. An evaluation of sub-
strate types in constructed wetland drainage treatment systems. In Mine Draina£ie
and Surface Mine Reclamation, Vol. 1: Mine Water and Mine Waste. U.S. Depart-
ment of the Interior, Pittsburgh, PA, pp. 389-398.

Chimney, M. J., and G. Goforth. 2006. History and description of the Everglades
Nutrient Removal Project, a subtropical constructed wetland in south Florida
(USA). Ecolo£tical En£!ineerin£ 27: 268-278.

Collins, J. N., and V. H. Resh. 1989. Guidelines for the ecological control of mosquitoes
in non-tidal wetlands of the San Erancisco Bay area. California Mosquito Vector
Control Association and University of California Mosquito Research Program,
Sacramento, CA.

Comin, F. A., J. A. Romero, V. Astorga, and C. Garcia. 1997. Nitrogen removal and
cycling in restored wetlands used as filters of nutrients for agricultural runoff.
Water Science and Technology 35: 255-261.

Conway, T. E., and J. M. Murtha. 1989. The Iselin Marsh Pond meadow. In D. A.
Hammer, ed.. Constructed Wetlands for Wastewater Treatment . Eewis Publishers,
Chelsea, Ml, pp. 139-144.

Cooke, J. G. 1992. Phosphorus removal processes in a wetland after a decade of receiv-
ing a sewage effluent. Journal of Environmental Quality 21: 733-739.

Cooper, P. F., and B. C. Findlater, eds. 1990. Constructed Wetlands in Water Pollution
Control. Pergamon Press, Oxford, UK. 605 pp.

Cronk, J. K. 1996. Constructed wetlands to treat wastewater from dairy and swine
operations: A review. A^friculture, Ecosystems and Environment 58: 97-114.

Davison, F., D. Pont, K. Bolton, and T. Headley. 2006. Dealing with nitrogen in
subtropical Australia: Seven case studies in the diffusion of ecotechnological inno-
vation. Ecolo£iical Engineering 2^-. 213-223.

Day, J. W., Jr., J. Ko, J., Rybczyk, D. Sabins, R. Bean, G. Berthelot, C. Brantley,
F. Cardoch, W. Conner, J. N. Day, A. J. Englande, S. Feagley, E. Hyfield, R.
Fane, J. Findsey, J. Mistich, E. Reyes, and R. Twilley. 2004. The use of wetlands
in the Mississippi delta for wastewater assimilation: A review. Ocean and Coastal
Mana£fement 47\ 671-691.

Day, J. W., Jr., R. Fane, M. Moerschbaecher, R. DeFaune, 1. Mendelssohn, J. Baus-
tian, and R. Twilley. 2013. Vegetation and soil dynamics of a Fouisiana estuary
receiving pulsed Mississippi River water following Hurricane Katrina. Estuaries
and Coasts 36: 665-682.

Deegan, F. A., D. S. Johnson, R. S.Warren, B. J. Peterson, J. W. Fleeger, S. Fagherazzi,
and W. M. Wollheim. 2012. Coastal eutrophication as a driver of salt marsh loss.
Nature 490: 388-392

694 Chapter 19 Wetlands and Water Quality

Dejong, J. 1976. The purification of wastewater with the aid of rush or reed ponds.
In J. Tourbier and R. W. Pierson, Jr., eds., Biolq/jical Control of Water Pollution.
University of Pennsylvania, Philadelphia, pp. 133-139.

Dierberg, F. E., and DeBusk, T. A. 2008. Particulate phosphorus transformations in
south Florida stormwater treatment areas used for Everglades protection. Ecolojj-
kal Enjjineerinjj 34, 100-115.

Entry, J. A., and A. Gottlieb. 2014. The impact of stormwater treatment areas and
agricultural best management practices on water quality in the Everglades Protec-
tion Area. Environmental Monitoring and Assessment 186, 1023-1037.

Ewel, K. C., and H. T. Odum, eds. 1984. Cypress Swamps. University Presses of
Florida, Gainesville. 472 pp.

Fennessy, M. S., and W. J. Mitsch. 1989. Treating coal mine drainage with an
artificial wetland. Research Journal of the Water Pollution Control Federation 61:
1691-1701.

Fink, D. F., and W. J. Mitsch. 2004. Seasonal and storm event nutrient removal by a
created wetland in an agricultural watershed. Ecolopfical Enpiineerinp; 2?>-. 313-325.

Fink, D. F., and W. J. Mitsch. 2007. Hydrology and nutrient biogeochemistry in a
created river diversion oxbow wetland. Ecological En^ineerinp;. 30: 93-162.

Gerheart, R. A. 1992. Use of constructed wetlands to treat domestic wastewater, city
of Areata, Galifornia. Water Science and Technology 26\ 1625-1637.

Gerheart, R. A., F. Klopp, and G. Allen. 1989. Gonstructed free surface wedands to
treat and receive wastewater: Pilot project to full scale. In D. A. Hammer, ed..
Constructed Wetlands for Wastewater Treatment. Lewis Publishers, Ghelsea, MI,
pp. 121-137.

Greenway, M. 2005. The role of constructed wetlands in secondary effluent treat-
ment and water reuse in subtropical and arid Australia. Ecological Enpfineerin^
25: 501-509.

Greenway, M., P. Dale, and H. Ghapman. 2003. An assessment of mosquito breeding
and control in four surface flow wedands in tropical-subtropical Australia. Water
Science & Technology 48: 249-256.

Gu, B., M. J. Ghimney, J. Newman, and M. K. Nungesser. 2006. Limnological char-
acteristics of a subtropical constructed wedand in south Florida (USA). Ecological
Enptineerinpt 27 ■. 345-360.

Headley, T. R., E. Herity, and L. Davison. 2005. Treatment at different depths and
vertical mixing within a 1-m deep horizontal subsurface-flow wedand. Ecological
Enptineerinpt 2S\ 567-582.

Hickman, S. G., and V. J. Mosca. 1991. Improving habitat quality for migratory
waterfowl and nesdng birds: Assessing the effeedveness of the Des Plaines River
wedands demonstration project. Technical Paper 1, Wetlands Research, Ghicago.
13 pp.

Hoagland, G. R., L. E. Gentry, M. B. David, and D. A. Kovacic. 2001. Plant nutrient
uptake and biomass accumulation in a constructed wedand. Journal of Freshwater
Ecology 16: 527-540

Jacks, G., A. Joelsson, and S. Fleischer. 1994. Nitrogen retendon in forested wedands.
Ambio2Z: 358-362.

References 695

Jackson, J. 1989. Man-made wetlands for wastewater treatment: Two case studies.
In D. A. Hammer, ed.. Constructed Wetlands for Wastewater Treatment . Lewis
Publishers, Chelsea, MI, pp. 574-580.

Juston, J., and DeBusk, T. A. 2006. Phosphorus mass load and outflow concentration
relationships in stormwater treatment areas for Everglades restoration. Ecolo£iical
En£tineerin£t 26, 206-223.

Juston, J., and DeBusk, T. A. 2011. Evidence and implications of the background
phosphorus concentration of submerged aquatic vegetation wetlands in stormwa-
ter treatment areas for Everglades restoration. Water Resources Research 41,
430-446.

Kadlec, R. H. 1999. Constructed wetlands for treating landfill leachate. In G. Mulam-
oottil, E. A. McBean, and F. Rovers, eds.. Constructed Wetlands for the Treatment
of Landfill Leachates . Lewis Publishers, Boca Raton, EL, pp. 17-31.

Kadlec, R. H. 2006. Free surface wetlands for phosphorus removal: The position of the
Everglades Nutrient Removal Project. 2006. Ecolo£iical En^ineerin£i 27 \ 361-379.

Kadlec, R. H. 2009a. Comparison of free water and horizontal subsurface treatment
wedands. Ecolo£iical En£iineerin£i 159-174.

Kadlec, R. H. 2009b. Wastewater treatment at the Houghton Lake wetland: Hydrol-
ogy and water quality. Ecological Engineering 35: 1287-1311.

Kadlec, R. H., and D. L. Hey. 1994. Constructed wetlands for river water quality
improvement. Water Science and Technolo£;y 29'. 159-168.

Kadlec, R. H., and R. L. Knight. 1996. Treatment Wetlands. CRC Press/Lewis Pub-
lishers, Boca Raton, EL. 893 pp.

Kadlec, R. H., and S. Wallace. 2009. Treatment Wetlands., 2nd ed. CRC Press, Boca
Raton, EL, 1016 pp.

Kearney, M. S., J. C. A. Riter, and R. E Turner. 2011. Freshwater river diver-
sions for marsh restoration in Louisiana: Twenty-six years of changing
vegetative cover and marsh area. Geophysical Research Letters 38, L16405.
doi : 1 0 . 1 029/20 1 1 GL047847 .

Knight, R. L. 1990. Wetland systems. In Natural Systems for Wastewater Treatment,
Manual of Practice ED- 16. Water Pollution Gontrol Federation, Alexandria, VA,

pp. 211-260.

Knight, R. L., T. W. McKim, and H. R. Kohl. 1987. Performance of a natural wet-
land treatment system for wastewater management. Journal of the Water Pollution
Control Eederation 59: 746-754.

Knight, R. L., V. W. E. Payne Jr., R. E. Borer, R. A. Glarke Jr., and J. H. Pries. 2000.
Gonstructed wedands for livestock wastewater management. Ecological En^ineer-
ing 15: 41-55.

Knight, R. L., W. E. Walton, G. F. O’Meara, W. K. Reisen, and R. Wass. 2003. Strate-
gies for effective mosquito control in constructed treatment wetlands. Ecolopfical
Engineering 21: 211-232.

Kovacic, D. A., M. B. David, L. E. Gentry, K. M. Starks, and R. A. Gooke. 2000. Effec-
tiveness of constructed wetlands in reducing nitrogen and phosphorus export
from agricultural tile drainage. /owma/ of Environmental Quality 29: 1262-1274.

696 Chapter 19 Wetlands and Water Quality

Lane, R. R., J. W. Day, D. Justic, E. Reyes, J. N. Day, and E. ELyfield. 2004. Changes
in stoichiometric Si, N and P ratios of Mississippi River water diverted through
coastal wetlands to the Gulf of Mexico. Estuarine Coastal & Shelf Science 60:
1-10.

Lane, R. R., J. W. Day Jr., and J. N. Day. 2006. Wetland surface elevation, vertical
accretion, and subsidence at three Louisiana estuaries receiving diverted Missis-
sippi River water. Wetlands 26\ 1130-1142.

Lane, R. R., J. W. Day Jr., B. D. Marx, E. Reyes, E. Hyfield, and J. N. Day. 2007.
The effects of riverine discharge on temperature, salinity, suspended sediment and
chlorophyll a in a Mississippi delta estuary measured using a flow-through system.
Estuarine Coastal & Shelf Science 74: 145-154.

Larson, A. C., L. E. Gentry, M. B. David, R. A. Cooke, and D. A. Kovacic. 2000.
The role of seepage in constructed wetlands receiving tile drainage. Ecolo£iical
Engineering 15: 91-104.

Leonardson, L., L. Bengtsson, T. Davidsson, T. Persson, and U. Emanuelsson. 1994.
Nitrogen retention in artificially flooded meadows. Ambio 23: 332-341.

Litchfield, D. K., and D. D. Schatz. 1989. Constructed wetlands for wastewater
treatment at Amoco Oil Company’s Mandan, North Dakota, refinery. In D. A.
Hammer, ed.. Constructed Wetlands for Wastewater Treatment. Lewis Publishers,
Chelsea, Ml, pp. 101-119.

Lundberg, C. J., R. R. Lane, and J. W. Day, Jr. 2014. Spatial and temporal variations in
nutrients and water-quality parameters in the Mississippi River-influenced Breton
Sound Estuary. Journal of Coastal Research 30: 328-336.

Mander, U, G. Dotro, Y. Ebie, S. Towprayoon, G. Ghiemchaisri, S. F. Nogueira, B.
Jamsranjav, K. Kasak, J. Truu, J. Tournebize, and W. J Mitsch. 2014. Green-
house gas emission in constructed wedands for wastewater treatment: a review.
Ecological En£!ineerin£! 66: 19-35.

Manyin, T., F. M. Williams, and L. R. Stark. 1997. Effects of iron concentration and
flow rate on treatment of coal mine drainage in wetland mesocosms: An exper-
imental approach to sizing of constructed wetlands. Ecological Engineering 9:
171-185.

Mitchell, D. S., A. J. Chick, and G. W. Rasin. 1995. The use of wetlands for water pol-
lution control in Australia: An ecological perspective. Water Science and Technology
32: 365-373.

Mitsch, W. J. 2012. What is ecological engineering.^ Ecological Engineering 45: 5-12.

Mitsch, W. J., J. K. Cronk, X. Wu, R. W. Nairn, and D. L. Hey. 1995. Phosphorus
retention in constructed freshwater riparian marshes. Ecological Applications 5:
830-845.

Mitsch, W. J., and K. M. Wise. 1998. Water quality, fate of metals, and predictive
model validation of a constructed wetland treating acid mine drainage. Water
Research 32: 1888-1900

Mitsch, W. J., X. Wu, R. W. Nairn, P. E. Weihe, N. Wang, R. Deal, and C. E.
Boucher. 1998. Creating and restoring wetlands: A whole-ecosystem experiment
in self-design. BioScience 48: 1019-1030.

References 697

Mitsch, W. J., A. J. Home, and R. W. Nairn. 2000. Nitrogen and phosphorus retention
in wetlands — Ecological approaches to solving excess nutrient problems. Ecologi-
cal Engineering 14: 1-7.

Mitsch, W. J., and S. E. Jorgensen. 2004. Ecolo£;ical En^jineerin^f and Ecosystem
Restoration. John Wiley & Sons, Hoboken, NJ. 411 pp.

Mitsch, W. J., N. Wang, L. Zhang, R. Deal, X. Wu, and A. Zuwerink. 2005a.
Using ecological indicators in a whole-ecosystem wetland experiment. In S. E.
Jorgensen, F-L. Xu, and R. Costanza, eds.. Handbook of Ecological Indicators for
Assessment of Ecosystem Health. CRC Press, Boca Raton, EL, pp. 211-235.

Mitsch, W. J., J. W. Day Jr., L. Zhang, and R. Lane. 2005b. Nitrate -nitrogen retention
by wetlands in the Mississippi River Basin. Ecological Engineering 24: 267-278.

Mitsch, W. J., L. Zhang, C. J. Anderson, A. Altor, and M. Hernandez. 2005c. Creating
riverine wetlands: Ecological succession, nutrient retention, and pulsing effects.
Ecolo£iical Engineering 25: 510-527.

Mitsch, W. J., L. Zhang, D. F. Fink, M. E. Hernandez, A. E. Altor, C. L. Tuttle
and A. M. Nahlik. 2008. Ecological engineering of floodplains. Ecohydrolo^y &
Hydrobiology 139-147.

Mitsch, W.J., L. Zhang, E. Waletzko, and B. Bernal. 2014. Validation of the ecosystem
services of created wedands: Two decades of plant succession, nutrient retention,
and carbon sequestration in experimental riverine marshes. Ecological Engineering
doi.org/10.1016/j.ecoleng. 2014.09. 108

Mitsch, W.J., L. Zhang, D. Marois, and K. Song. 2015. Protecting the Florida Ever-
glades wetlands with wetlands: Can stormwater phosphorus be reduced to olig-
otrophic conditions? Ecological Engineering., in press.

Morris, J. T., G. P. Shaffer, and J. A. Nyman. 2013. Brinson Review: Perspectives on
the influence of nutrients on the sustainability of coastal wetlands. Wetlands 33:
975-988.

Moustafa, M. Z. 1999. Nutrient retention dynamics of the Everglades nutrient
removal project. Wetlands 19: 689-704.

Moustafa, M. Z., M. J. Chimney, T. D. Fontaine, G. Shih, and S. Davis. 1996. The
response of a freshwater wetland to long-term “low level” nutrient loads — marsh
efficiency. Ecological Engineering 7 ■. 15-33.

Mulamoottil, G., E. A. McBean, and F. Rovers, eds. 1999. Constructed Wetlands for
the Treatment of Landfill Leachates . Lewis Publishers, Boca Raton, EL. 281 pp.

Nairn, R. W., and W. J. Mitsch. 2000. Phosphorus removal in created wetland ponds
receiving river overflow. Ecological Engineering 14: 107-126.

Newman, J. M., J. C. Clausen, and J. A. Neafsey. 2000. Seasonal performance of a wet-
land constructed to process dairy milkhouse wastewater in Connecticut. Ecological
Engineering 14: 181-198.

698 Chapter 19 Wetlands and Water Quality

Newman, S., and K. Pietro. 2001. Phosphorus storage and release in response to
flooding: Implications for Everglades stormwater treatment areas. Ecolo£iical En£ii-
neering 18: 23-38.

Nguyen, L. M. 2000. Phosphate incorporation and transformation in surface sedi-
ments of a sewage-impacted wetland as influenced by sediment sites, sediment
pH and added phosphate concentration. Ecological Engineering 14: 139-155.

Nguyen, L. M., J. G. Cooke, and G. B. McBride. 1997. Phosphorus retention and
release characteristics of sewage-impacted wetland sediments. Water, Air, and Soil
Pollution 100: 163-179.

Niswander, S. F., and W. J. Mitsch. 1995. Functional analysis of a two-year-old created
in-stream wetland: Hydrology, phosphorus retention, and vegetation survival and
growth. Wetlands 15: 212-225.

Odum, H. T., K. C. Ewel, W. J. Mitsch, and J. W. Ordway. 1977. Recycling treated
sewage through cypress wetlands in Florida. In F. M. D’ltri, ed.. Wastewater Ren-
ovation and Reuse. Marcel Dekker, New York, pp. 35-67.

Ogden, M. H. 1999. Constructed wetlands for small community wastewater treatment .
Paper presented at Wedands for Wastewater Recycling Conference, November 3,
1999, Baltimore. Environmental Concern, St. Michaels, MD.

Paudel, R., J.-K. Min, and J. W. Jawitz. 2010. Management scenario evaluation for a
large treatment wetland using a spatio-temporal phosphorus transport and cycling
model. Ecological Engineering; 36: 1627-1638.

Paudel, R., and J. W. Jawitz. 2012 . Does increased model complexity improve descrip-
tion of phosphorus dynamics in a large treatment wetland? Ecolo£;ical En£;ineerin£i
42: 283-294.

Phipps, R. G., and W. G. Crumpton. 1994. Factors affecting nitrogen loss in
experimental wetlands with different hydrologic loads. Ecological En£;ineerin^ 3:
399-408.

Pietro, K. 2012. Synopsis of the Everglades stormwater treatment areas, water year
1996-2012. Technical Publication ASB-WQTT-12-001, South Florida Water
Management District, West Palm Beach, FT. 56 pp.

Raisin, G. W, and D. S. Mitchell. 1995. The use of wetlands for the control of
non-point source pollution. Water Science and Technology 32: 177-186.

Raisin, G. W, D. S. Mitchell, and R. F. Croome. 1997. The effectiveness of a small
constructed wetland in ameliorating diffuse nutrient loadings from an Australian
rural catchment. Ecolo£;ical Engineering 9 \ 19-35.

Reddy, K. R., R. H. Kadlec, M. J. Chimney, and W. J. Mitsch, eds. 2006. The
Everglades Nutrient Removal Project. Special issue of Ecolopfical Engineering 27:
265-379.

Reinelt, F. E., and R. R. Horner. 1995. Pollutant removal from stormwater runoff
by palustrine wetlands based on comprehensive budgets. Ecolopfical Engineering
4: 77-97.

Richardson, C. J., and C. B. Craft. 1993. Effective phosphorus retention in
wetlands — Fact or Action? In G. A. Moshiri, ed.. Constructed Wetlands for Water
Quality Improvement. CRC Press, Boca Raton, FT, pp. 271-282.

References 699

Richardson, C. J., S. Qian, C. B. Craft, and R. G. Qualls. 1997. Predictive mod-
els for phosphorus retention in wedands. Wetlands Ecology and Management 4:
159-175.

Rivera-Monroy, V. H., B. Branoff, E. Meselhe, A. McCorquodale, M. Dortch, G.
D. Steyer, J. Visser, and H. Wang. 2013. Landscape-level estimation of nitrogen
removal in coastal Louisiana wetlands: Potential sinks under different restoration
scenarios. Journal of Coastal Research 67: 75-87.

Sartoris, J. J., J. S. Thullen, L. B. Barber, and D. E. Salas. 2000. Investigation of nitro-
gen transformations in a southern Galifornia constructed wastewater treatment
wetland. Ecologtical Engtineering; 14: 49-65.

Schaafsma, J. A., A. H. Baldwin, and G. A. Streb. 2000. An evaluation of a constructed
wetland to treat wastewater from a dairy farm in Maryland, USA. Ecological Engi-
neering 14: 199-206.

Seidel, K. 1964. Abbau von Bacterium GoU durch hohere Wasserpflanzen. Naturwis-
senschaften 51: 395.

Seidel, K. 1966. Reinigung von Gewassern durch hohere Pflanzen. Naturwis-
senschaften 53: 289-297.

Seidel, K., and H. Happl. 1981. Pflanzenldaranlage “Krefelder system.” Sicherheit in
Chemic und Umbelt 1: 127-129.

Solano, M. L., P. Soriano, and M. P. Giria. 2004. Gonstructed wetlands as a sustainable
solution for wastewater treatment in small villages. Biosystems Engineering 87:
109-118.

Spieles, D. J., and W. J. Mitsch. 2000a. The effects of season and hydrologic and
chemical loading on nitrate retention in constructed wetlands: A comparison of
low and high nutrient riverine systems. Ecologfical Engfineerin^ 14: 77-91.

Spieles, D. J., and W. J. Mitsch. 2000b. Macroinvertebrate community structure in
high- and low-nutrient constructed wetlands. Wetlands 20: 716-729.

Stark, L. R., and L. M. Williams. 1995. Assessing the performance indices and design
parameters of treatment wedands for H, Le, and Mn retention. Ecological Engfi-
neering 5: 433-444.

Steiner, G. R., J. T. Watson, D. Hammer, and D. L. Marker, Jr. 1987. Municipal
wastewater treatment with artificial wetlands — A TVA/Kentucky demonstration.
In K. R. Reddy and W. H. Smith, eds.. Aquatic Plants for Wastewater Treatment
and Resource Recovery. Magnolia Publishing, Orlando, PL, p. 923.

Steiner, G. R., and R. J. Preeman, Jr. 1989. Gonfiguration and substrate design con-
siderations for constructed wetlands for wastewater treatment. In D. A. Hammer,
ed.. Constructed Wetlands for Wastewater Treatment . Lewis Publishers, Ghelsea,
MI,pp. 363-378.

Tanner, G. G. 1996. Plants for constructed wetland treatment systems — A comparison
of the growth and nutrient uptake of eight emergent species. Ecological Engineer-
ing 7: 59-83.

Tanner, G. G., J. S. Glayton, and M. P. Upsdell. 1995. Effect ofloading rate and plant-
ing on treatment of dairy farm wastewaters in constructed wetlands. II. Removal
of nitrogen and phosphorus. Water Research 29: 27-34.

700 Chapter 19 Wetlands and Water Quality

Tarutis, W. J., L. R. Stark, and F. M. Williams. 1999. Sizing and performance estima-
tion of coal mine drainage wetlands. Ecological Engineering 12: 353-372.

Teal, J. M., R. Best, J. Caftfey, C. S. Hopkinson, K. L. McKee, J. T. Morris, S. New-
man, and B. Orem. 2012. Mississippi River freshwater diversions in southern
Louisiana: Effects on wedand vegetation, soils, and elevation. Final Report to the
State of Louisiana and the U.S. Army Corps of Engineers through the Louisiana
Coastal Area Science & Technology Program and National Oceanic and Atmo-
spheric Administration. 49 pp.

Thullen, J. S., J. J. Sartoris, and S. M. Nelson. 2005. Managing vegetation in
surface-flow wastewater-treatment wetlands for optimal treatment performance.
Ecological En£iineerin£i 25: 583-593.

Teiter, S., and U. Mander. 2005. Emission of N2O, N2, CH4, and CO2 from con-
structed wetlands for wastewater treatment and from riparian buffer zones. Eco-
lo£iical Engineering 2S\ 528-541

Vymazal, J. 1995. Constructed wetlands for wastewater treatment in the Czech
Republic — state of the art. Water Science and Technology 32: 357-364.

Vymazal, J. 1998. Czech Republic. In J. Vymazal, H. Brix, P. F. Cooper, M. B. Green,
and R. Haberl, eds.. Constructed Wetlands for Wastewater Treatment in Europe.
Backhuys Publishers, Leiden, the Netherlands, pp. 95-121.

Vymazal, J. 2002. The use of sub-surface constructed wetlands for wastewater treat-
ment in the Czech Republic: 10 years experience. Ecolopfical Enpfineerinpi 18:
633-646.

Vymazal, J., ed. 2005. Constructed wedands for wastewater treatment. Special Issue
of Ecological En^ineerinp; 25: 475-621.

Vymazal, J. 2013. Emergent plants used in free water surface constructed wedands:
A review. Ecolojjical Enjjineerinpf 61P\ 582-592.

Vymazal, J., and L. Kropfelova. 2005. Growth of Phrapfmites australis and Phalaris
arundinacea in constructed wetlands for wastewater treatment in the Gzech
Republic. Ecological Engineering 2S\ 606-621.

Weller, M. W. 1994. Freshwater Marshes., 3rd ed. University of Minnesota Press, Min-
neapolis. 192 pp.

White, J. S., S. E. Bayley, and P. J. Gurds. 2000. Sediment storage of phosphorus
in a northern prairie wedand receiving municipal and agro-industrial wastewater.
Ecological Enptineerinpt 14: 127-138.

Wieder, R. K., and G. E. Lang. 1984. Influence ofwedands and coal mining on stream
water chemistry. Water, Air, and Soil Pollution 23: 381-396.

Wilhelm, M., S. R. Lawry, and D. D. Hardy. 1989. Greadon and management of wet-
lands using municipal wastewater in northern Arizona: A status report. In D. A.
Hammer, ed.. Constructed Wetlands for Wastewater Treatment. Lewis Publishers,
Ghelsea, Ml, pp. 179-185.

Appendix A

Wetland Losses by State in the United
States, 1780s-1980s

State

Original Wetlands
Circa 1780
(xl,000 ha)

National Wetlands
Inventory, mid-1980s
(xl,000 ha)

Change

(percent)

Alabama

3,063

1,531

-50

Alaska

68,799

-0.1

Arizona

377

243

-36

Arkansas

3,986

1,119

-72

California

2,024

184

-91

Colorado

809

405

-50

Connecticut

271

-74

Delaware

194

-54

Florida

8,225

4,467

-46

Georgia

2,769

2,144

-23

Hawaii

-12

Idaho

355

156

-56

Illinois

3,323

508

-85

Indiana

2,266

304

-87

Iowa

1,620

171

-89

Kansas

340

176

-48

Kentucky

634

121

-81

Louisiana

6,554

3,555

Maine

2,614

2,104

-19

Maryland

668

178

-73

Massachusetts

331

238

-28

Michigan

4,533

2,259

Minnesota

6,100

3,521

-42

Mississippi

3,995

1,646

Missouri

1,960

260

-87

{continued)

701

702 Appendix A Wetland Losses by State in the United States, 1780s-1980s

state

Original Wetlands
Circa 1780
(xl.OOO ha)

National Wetlands
Inventory, mid-1980s
(xl,000 ha)

Change

(percent)

Montana

464

340

-27

Nebraska

1,178

771

-35

Nevada

197

-52

New Hampshire

-9

New Jersey

607

370

-39

New Mexico

291

195

-33

New York

1,037

415

-60

North Carolina

4,488

2,300

-44

North Dakota

1,994

1,008

-49

Ohio

2,024

195

-90

Oklahoma

1,150

384

-67

Oregon

915

564

-38

Pennsylvania

456

202

-56

Rhode Island

-37

South Carolina

2,596

1,885

-27

South Dakota

1,107

720

-35

Tennessee

784

318

-59

Texas

6,475

3,080

-52

Utah

325

226

-30

Vermont

138

-35

Virginia

748

435

-42

Washington

546

380

-31

West Virginia

-24

Wisconsin

3,966

2,157

-46

Wyoming

809

506

-38

Total wetlands

158,395

111,060

-30

Total “lower 48”

89,491

42,240

-53

Source'. Dahl, T. E. 1990. Wetlands losses in the United States, 1780s to 1980s. U.S. Department of Interior,
Fish and Wildlife Service, Washington, DC. 21 pp.

Appendix B

Useful Wetland Web Pages

International

U.S. National Ramsar Committee, http://usnrc.net
The Ramsar Convention on Wetlands, www.ramsar.org
Wetlands International, www.wetlands.org
International Peat Society, www.peatsociety.org

International Association of Ecology (INTECOL), www.intecol.net/pages/index.php
United Nations Environment Programme, www.unep.org
IPCC, Intergovernmental Panel on Climate Change, www.ipcc.ch
The International Union for Conservation of Nature (lUCN), www.iucn.org

Wetland Restoration

The Marshlands of Mesopotamia, http://whc.unesco.org/en/tentativelists/1838/ The Bois-des-Bel exper-
imental peatlands, Quebec, Canada www.gret-perg.ulaval.ca/recherche/themes-de-recherche/diversite-
floristique/bois-des-bel

The Skjern River, Denmark www.globalrestorationnetwork.org/database/case-study/?id=115
The Comprehensive Everglades Restoration Plan (CREP) www.nps.gov/ever/naturescience/cerp.htm; www
. evergladesplan . org

Florida Uniform Mitigation Assessment Method (UMAM) www.dep.state.fl.us/water/wetlands/mitigation/
umam http : / / sfrc.ufl.edu / ecohydrology/UMAM_Training_Manual_ppt.pdf
Ohio Rapid Assessment Method (ORAM), www.epa.state.oh.us/dsw/401/ecology.aspx
State of Washington Wetland Rating System http:/ /www.ecy.wa.gov/programs/sea/wetlands/ratingsystems/
2014updates.html

University Wetland Programs

Everglades Wetland Research Park, FGCU, www.fgcu.edu/swamp
Duke University Wetland Center, http:/ /nicholas. duke.edu/wetland

703

704 Appendix B Useful Wetland Web Pages

Howard T. Odum Center for Wedands, University of Florida, http://cfw.essie.ufl.edu
LSU School of the Coast and Environment, www.sce.lsu.edu/about

U.S. Government

uses National Wetlands Research Center, www.nwrc.usgs.gov
US EPA Wetlands, http://water.epa.gov/type/wetlands
USDA/NRCS Wetlands

General: www.nrcs.usda.gov/wps/portal/nrcs/main/national/water/wetlands
Plants: https://plants.usda.gov/core/wetlandSearch
Soils: www.nrcs.usda.gov/wps/portal/nrcs/main/soils/use/hydric
U.S. Fish and Wildlife Service National Wetlands Inventory, www.fws.gov/wetlands

U.S. Army Corps of Engineers, www.usace.army.mil/Missions/CivilWorks/RegulatoryProgramandPermits
.aspx

Societies /N G Os

Ducks Unlimited, www.ducks.org

Society of Wetland Scientists, www.sws.org

Association of State Wetland Managers, www.aswm.org

The Nature Conservancy, www.nature.org

Society for Ecological Restoration, www.ser.org

American Ecological Engineering Society, www.ecoeng.org

National Audubon Society, www.audubon.org

Scientific Journals

Ecological Engineering, www.journals.elsevier.com/ecologicaTengineering
Wetlands, www. sws .org/Publications/ wetlands - j ournal . h tml
Wetlands Ecology and Management, http:/ /link.springer.com/journal/11273
Estuaries and Coasts, http://link.springer.com/journal/12237

Adventure /Entertainment

“Bill and Ulo’s Excellent Adventure” (in the Okavango Delta, Botswana), www.youtube.com/watch.>v=
ORg97COzuRM

Mega Python vs. Gatoroid trailer, www.youtube.com/watch?v=S8pirKwkxbO

Swamp Thing trailer, YouTube video. Image Entertainment, Inc., uploaded on Apr 12, 2011, www.youtube
.com/watch.>v=kzbqK4nw3R8

“Olentangy River Wetland Research Park,” YouTube video, WOSU Public Media (www.wosu.org), uploaded
December 2, 2008, www.youtube. com/watch?v=KL-34AZPprE
Wetland and Carbon lecture — Bill Mitsch, Montpellier France, October 25, 2013, YouTube video, Everglades
Wetiand Research Park, uploaded December 4, 2013, www. youtube. com/watch?v=AaCUUSOglxs

“Wetlands” by Bill Mitsch, Bonita Springs, Elorida, April 4, 2013. YouTube Video Produced by Charlotte Har-
bor National Estuarine Program Office, www.youtube. com/watch.>v=MYZGXXwlIOo&feature=youtu. be.

Appendix C

Useful Conversion Factors

Multiply

To Obtain

LENGTH

centimeters (cm)

0.3937

inches

feet

0.3048

meters (m)

inches

2.54

centimeters (cm)

kilometers (km)

0.6214

miles

meters (m)

3.2808

feet

meters (m)

39.37

inches

meters (m)

1.0936

yards

miles

1.6093

kilometers (km)

yards

0.9144

meters (m)

AREA

acres

0.4047

hectares (ha)

2.47

acres

hectares (ha)

10,000

square meters (m^)

acres

4047

square meters (m^)

hectares (ha)

0.01

square kilometers (km^)

100

hectares (ha)

square meters (m^)

0.0001

hectares (ha)

VOLUME

cubic feet

0.02834

cubic meters (m^)

35.31

cubic feet

cubic centimeters (cm^)

10-3

liters (L)

acre-feet

1223.5

cubic meters (m^)

gallons

3.785

liters (L)

gallons

0.003785

cubic meters (m^)

264.2

gallons

liters (L)

0.2642

gallons

{continued)

70S

706 Appendix C Useful Conversion Factors

Multipiy

To Obtain

FLOW

cubic feet per second (cfs)

0.002832

cubic meters per second (m® s-^)

cubic feet per second (cfs)

10.1952

cubic meters per hour (m® hr-^)

cubic feet per second (cfs)

448.86

gallons per minute (gpm)

cubic meters per second (m^ s“^)

35.31

cubic feet per second (cfs)

cubic meters per second (m^ s“^)

3600

cubic meters per hour (m® hr-^)

gallons per minute (gpm)

0.002228

cubic feet per second (cfs)

gallons per minute (gpm)

0.06308

liters per second (L s-^)

gallons per minute (gpm)

0.00379

cubic meters per minute
(m® min-4)

MASS

grams (g)

0.002205

pounds

grams (g)

0.001

kilograms (kg)

2.2046

pounds

kilograms (kg)

1000

grams (g)

pounds

453.6

grams (g)

pounds

0.4536

kilograms (kg)

metric tons (t)

2205

pounds

metric tons (t)

1000

kilograms (kg)

PRESSURE

atmosphere (atm)

1.01325x105

pascal (Pa)

atmosphere (atm)

760

millimeters of mercury (mm Hg)

FLUX OF MASS

grams per square meter per year
(gm-2 yr-i)

kilograms per hectare per year
(kg ha-4 y|.-ij

grams per square meter per year
(gm-2 yr-i)

8.924

pounds per acre per year

kilograms per hectare per year
(kgha“^ yr“^)

0.1

grams per square meter per year
(gm-2yr-i)

pound per acre per year

1.12

kilograms per hectare per year
(kg ha-4 y|.-ij

pounds per acre per year

0.112

grams per square meter per year
(gm-2yr-4)

ENERGY

British thermai units (BTU)

0.2530

kilocalories (kcal)

British thermai units (BTU)

1054

joules (J)

calories (cal)

4.1869

joules (J)

calories (cal)

0.001

kilocalories (kcal)

joules (j)

0.239

calories (cal)

joules (J)

2.390x10-4

kilocalories (kcal)

1000

calories (cal)

kilocalories (kcal)

3.968

British thermal units (BTU)

kilocalories (kcal)

4183

joules (J)

kilocalories (kcal)

4.183

kilojoules (kJ)

kilocalories (kcal)

0.001162

kilowatt-hours (kWhr)

kilojoule (kJ)

0.239

kilocalories (kcal)

kilowatt-hours (kWhr)

860.5

kilocalories (kcal)

kilowatt-hours (kWhr)

3.6 X 10®

joules (J)

Appendix C Useful Conversion Factors 707

Multiply

To Obtain

langley (ly)^

calories per square centimeter (cal cm-^)

langley (ly)

kilocalories per square meter (kcal m-^)

POWER

horsepower

0.7457

kilowatts (kW)

horsepower

10.70

kilocalories per minute (kcal/min)

kllocalories/day (kcal/day)

6.4937 X 10-5

horsepower

kllocalories/day (kcal/day)

4.8417 X 10-5

kilowatts (kW)

1.341

horsepower

kilowatts (kW)

14.34

kilocalories per minute (kcal/min)

kilowatts (kW)

1000

watts (W)

watt (W)

joule per second (J/sec)

PRIMARY PRODUCTIVITY/ENERGY FLOW

grams dry weight (g-dw)

4.5

kilocalories (kcal)

grams dry weight (g-dw)

0.45

grams C (g-C)

grams 0^ (g-0^)^

3.7

kilocalories (kcal)

grams Oj (g-Oj)^

0.375

grams C (g-C)

grams C (g-C)^

kilocalories (kcal)

grams C (g-C)^

2.67

grams Oj (g-Oj)

kilocalories (kcal)

4.18

kilojoules (kJ)

kilocalories (kcal)^

0.1

grams C (g-C)

STOICHIOMETRY OF ORGANIC MATTER^

molar ratio

106C:16N:1P

weight ratio

41C:7.2N:1P

CONCENTRATIONS IN WATER

part per thousand (ppt)

grams per liter (gf-^)

part per million (ppm)

milligrams per liter (mgL-^)

parts per million (ppm)

grams per cubic meter (gm-^)

parts per million (ppm)

1000

parts per billion (ppb)

micrograms per liter (pg L-^)

milligrams per liter (mgL“^)

1000

micrograms per liter (pg L-^)

millimolarity (m mole/L)

molecular weight

milligrams per liter (mgf-^)

micromolarity (p mole/L)

molecular weight

micrograms per liter (pg L-^)

microgram-atoms per liter
(pg-atom L“^)

molecular weight

micrograms per liter (pg L-^)

milligrams per liter (mgL“^)

ionic charge/
molecular weight

milliequivatents per liter (meqL-^)

milliequivatents per liter (meq L“^)

molecular

weight/ionic

charge

milligrams per liter (mg L-^)

micromhos per cm (pmho cm“^)

microSiemens per centimeter (pScm-^)

^Solar constant = radiant energy at outer limit of earth’s atmosphere ~2.0 langleys per min (ly min ^ ).
^Based on general photosynthetic equation showing the production of glucose:
eCOj +12H20 + (118 X 6kcal)^C^Hj20^ +6O2 d-aHjO.

^Based on Redfield (1958) of plankton organic matter {CH20)jq^(NH3)j^(H2P0^).

Reference

Redfield, A. C. 1958. The biological control of chemical factors in the environment.
American Scientist 46: 206-226.

Glossary

Aapa peatlands — ^Also called string bogs and patterned fens; peatlands identified by watertracks of long, narrow
alignment of the high peat hummocks (strings) that form ridges perpendicular to the slope of the peatland
and are separated by deep pools (flarks).

ADH — ^Alcohol dehydrogenase, the enzyme that catalyzes the reduction of acetaldehyde to ethanol in fermen-
tation.

Adventitious roots — Roots that develop from some part of a vascular plant other than the seed. Usually they
originate from the stem, and while common in most plants, they also develop as adaptations to anoxia in
both flood-tolerant trees (e.g., ScilixznA Alnus) and herbaceous species, and flood-intolerant (e.g., tomato)
plants just above the anaerobic zone when these plants are flooded.

Aerenehyma — Large air spaces in roots and stems of some wetland plants that allow the diffusion of oxygen
from the aerial portions of the plant into the roots.

Alcohol dehydrogenase — See ADH.

Allochthonous — Pertains to material that is imported into an ecological system of interest from outside that
system; usually refers to organic material and/or nutrients and minerals.

Allogenic succession — Ecosystem development whereby the distribution of species is governed by their indi-
vidual responses to their environment with little or no feedback from organisms to their environment. Also
called individualistic hypothesis, continuum concept, and Gleasonian succession.

Alluvial plain — The floodplain of a river, where the soils are alluvial deposits carried in by the overflowing river.

Ammonia volatilization — NHj released to the atmosphere as a gas.

Anadromous — Refers to marine species that spawn in freshwater streams.

Anammox — ^Abbreviation for anaerobic ammonium oxidation that leads to conversion of nitrite-nitrogen to
dinitrogen gas.

Anaerobic — Refers to oxygenless conditions.

Anoxia — ^Waters or soils with no dissolved oxygen.

Artificial wetland — See Constructed wetland.

Aspect ratio — length to width ratio of a constructed wetland. Recommended to be 10:1 or higher for
flow-through treatment wetlands.

Assimilatory nitrate reduction — Nitrate (NO3) assimilated by plants or microbes and converted into biomass.

Assimilatory sulfate reduction — Process in the sulfur cycle whereby sulfur-reducing obligate anaerobes such
as Desulfovibrio bacteria utilize sulfates as terminal electron acceptors in anaerobic respiration.

708

Glossary 709

Autochthonous — Pertains to material that is produced within the ecological system of interest (e.g., organic
material produced by photosynthesis). See Allochthonous.

Autogenic succession — Clementian theory of succession of ecosystems whereby vegetation occurs in recog-
nizable and characteristic communities; community change through time is brought about by the biota;
changes are linear and directed toward a mature stable climax ecosystem.

Bankfull discharge — Streamflow at which a river begins to overflow onto its floodplain.

Basin wetland — ^A wetland that is hydrologically isolated with little or no flooding from streams, rivers, or
tides.

Billabong — ^Australian term for a riparian wetland that is periodically flooded by the adjacent stream or river.

Biogeochemical cycling — The transport and transformation of chemicals in ecosystems.

Blanket bogs — In humid climates, peat that blankets large areas far from the site of the original peat accumu-
lation, through the process of paludification.

BOD — Biochemical oxygen demand, a biological test for degradable organic matter in water.

Bog — ^A peat- accumulating wetland that has no significant inflows or outflows and supports acidophilic mosses,
particularly Sphcignum.

Bottomland — Lowland along streams and rivers, usually on alluvial floodplains, that is periodically flooded.

Bottomland hardwood forest — Term used principally in southeastern and eastern United States to mean a
mesic riparian forested ecosystem along a higher order stream or river that is subject to intermittent to
frequent flooding from that stream or river; dominated by oaks and other deciduous hardwood tree species.

Bidk density — Dry weight of a known volume of soil, divided by that volume.

Buttress — Swollen bases of tree trunks growing in water.

Cajun — Term used for culture of former French-speaking immigrants who have lived for several centuries in
the swamps of the Louisiana delta.

Carbon sequestration — The permanent retention of carbon in an ecosystem, usually in its soil. See also Seques-
tration.

Carr — Term used in Europe for forested wetlands characterized by alders {Alnus) and willows (Salix).

Cat clays — When coastal wetlands are drained, soil sulfides often oxidize to sulfuric acid, and the soils become
too acidic to support plant growth.

Cation exchange capacity — The sum of exchangeable cations (positive ions) that a soil can hold.

Cheia — ^Annual period of flooding from March through May in the Pantanal region of South America that
supports luxurious aquatic plant and animal life. See also Enchente, Seca, and Vazante.

Clay depletions — Clay is selectively removed along root channels after iron and manganese oxides have been
depleted in wetland soils, only to redeposit as clay-coatings on soil particles below the clay depletions.

Coastal squeeze — ^A concept related to sea level change where coastal wetlands can be “squeezed out” during
sea level rise if there are human-constructed impediments that restrict the wetlands from moving inland.

Conforms — ^A quantitative biological test for the presence of colon bacteria or related forms; because of their
ubiquitous presence, they are used as a presumptive index of general bacterial contamination of water.

Concentric domed bog — ^A concentric pattern of pools and peat communities formed around the most ele-
vated part of a bog.

Constructed wetland — ^A wetland developed on a former uplands to create poorly drained soils and wetland
flora and fauna for the primary purpose of contaminant or pollution removal from wastewater or runoff.
See also Treatment wetland

Consiuner surplus — In economics, the net benefit of a good to the consumer.

Continuum concept — See Allogenic succession.

Coprecipitation of phosphorus — Some calcium phosphate is precipitated along with the major precipitation
of calcium carbonate in alkaline waters.

Created wetland — ^A wetland constructed where one did not exist before.

Cmnbungi swamp — Cattail ( Typha) marsh in Australia.

710 Glossary

Cumulative loss — When ecosystems such as wetlands are lost, usually as a result of human development, one
small piece at a time, with the cumulative loss being substantial.

Cypress domes — Also called cypress ponds or cypress heads; poorly drained to permanently wet depressions
dominated by pond cypress {Taxodium distichum van nutans). Called domes because the cypress grows
more vigorously in the center than around the perimeter of the dome, giving it a domed appearance from
a distance.

Cypress strand — K diffuse freshwater stream flowing through a shallow forested depression (dominated by
Taxodium) on a gently sloping plain.

Dabbling duck — ^Waterfowl mainly in the family Anatidae (swans, geese and ducks) that feed mainly at the
surface rather than by diving.

Dalton’s law — Flux is proportional to a pressure gradient. An example of a process that follows Dalton’s law
is evaporation, which is proportional to the difference between the vapor pressure at the water surface and
the vapor pressure in the overlying air.

Dambo — ^A seasonally waterlogged and grass-covered linear depressions in headwater zone of rivers with no
marked stream channel or woodland vegetation. Term is ChiChewa (Central Africa) dialect meaning
“meadow grazing.”

Darcy’s law — Groundwater equation that states that flow of groundwater is proportional to a hydraulic gradi-
ent and the hydraulic conductivity, or permeability, of the soil or substrate.

Delineation — Technique of determining an exact boundary of a wetland. Used for identifying jurisdictional
wetlands in United States.

Delta — Location where rivers meet the sea and deposit sediments, often in a broad alluvial fan; there are also
examples of inland deltas such as the Peace -Athabasca Delta in Canada and the Okavango Delta in Botswana
where the water never reaches the sea.

Demand curve — Economist’s estimate of consumer benefits.

Denitrification — Process in the nitrogen cycle carried out by microorganisms in anaerobic conditions, where
nitrate acts as a terminal electron acceptor, resulting in the loss of nitrogen as it is converted to nitrous
oxide (N2O) and nitrogen gas (Nj).

Designer wetland — Created or restored wetland in which certain plant species or other organisms are intro-
duced and the success or failure of those plants or organisms is used as the indicator of success or failure
of that wetland.

Detention time — ^A measure of the length of time a parcel of water stays in a wetland; equivalent to the turnover
time or retention time and the inverse of the turnover rate. Detention time is the term used most frequently
for designing treatment wetlands. See also Retention time.

Discharge wetland — ^Wetland that has surface water (or groundwater) level lower hydrologically than the sur-
rounding water table, leading to an inflow of groundwater.

Dissimilatory nitrate reduction to ammonia (DNRA) — Conversion of nitrate -nitrogen to ammonium-
nitrogen.

Dissimilatory nitrogen reduction — Several pathways of nitrate reduction, particularly nitrate reduction to
ammonia and denitrification. It is called dissimilatory as the nitrogen is not assimilated into a biological
cell.

Diversion wetland — ^Wetland created or enhanced by diversion of an adjacent body of water, usually a river.
Created diversion wetlands along rivers in upper watersheds are similar to oxbows or billabongs. River
diversions in deltas are means to re-establish deltaic distributaries.

Diving duck — Ducks that feed by diving beneath the surface of the water. They are in a distinct subfam-
ily, Aythyinae, of the large Anatidae family that includes ducks, geese, and swans. Also commonly called
pochards or scaups.

DMS — Dimethyl sulfide, one of the gases given off by wetlands.

Drop roots — See Prop roots.

Glossary 711

Duck stamps — Stamps sold in several countries to hunters to help pay for the protection of waterfowl habitat.
The Duck Stamp program in the United States started in 1934.

Eat-out — major wetland vegetation removal by herbivory, often by geese or muskrats.

Ebullitive flux — Flux of gases from wetland soils as bubbles or diffusion to the surface of the water and then
to the atmosphere.

Ecological engineering — The design, creation, and restoration of ecosystems for the benefit of humans and
nature.

Ecosystem engineers — Plants, animals, and microbes that carry out essential biological feedbacks in ecosys-
tems. Examples in wetlands are beavers and muskrats.

Ecosystem services — values that ecosystems provide to humans; similar to ecosystem values. Has been divided
into three categories related to human well-being: provisioning, regulating, and cultural.

Embodied energy — The total energy required to produce a commodity.

Emergy — Calculation of total energy requirement for any product in nature or humanity based on using trans-
formities. Short tor “energy memory.” See H. T. Odum (1996).

Enchente — Period of rising waters from December through February in Pantanal region of South America. See
also Cheia, Seca, and Vanzante.

Ericaceous plants — Flowering plants of the family Ericaceae, which, as a group, are acid-loving or acid-tolerant
plants that often dominate bogs and other sites with acidic substrates.

Estuary — General location where rivers meet the sea and freshwater mixes with saltwater.

Eutrophic — Nutrient rich; generally used in lake classification, but is also applicable to peatlands.

Eutrophication — Process of aquatic ecosystem development whereby an ecosystem such as a lake, estuary, or
wetland goes from an oligotrophic (nutrient poor) to eutrophic (nutrient rich) condition. If caused by
humans, it is called cultural eutrophication.

Excentric raised bogs — Bogs that form from previously separate basins on sloping land and form elongated
hummocks and pools aligned perpendicular to the slope.

Eacultative — ^Adapted equally to either wet or dry condition. Usually used in the context of vegetation adapted
to growing in saturated soils or upland soils.

Een — ^A peat-accumulating wetland that receives some drainage from surrounding mineral soil and usually
supports marshlike vegetation.

Eermentation — Partial oxidation of organic matter, when organic matter itself is the terminal electron acceptor
in anaerobic respiration by microorganisms; forms various low-molecular-weight acids and alcohols and
COj. Also called glycolysis.

Eibrists — See Peat.

Elarks — See Aapa peatlands.

Elood duration — The amount of time that a wetland is in standing water.

Elood frequency — The average number of times that a wetland is flooded during a given period.

Elood peak — Peak runoff into a wetland caused by a specific rainfall event.

Elood pulse concept (EEC) — Pulsing river discharge as the major force controlling biota in river floodplains,
including the lateral exchange between floodplains and river channels.

Eluted trunk — Flared tree trunks at the ground surface that occurs on some trees growing in wet conditions.

Eolists — Organic soils caused by excessive moisture (precipitation > evapotranspiration) that accumulate in
tropical and boreal mountains; these soils are not classified as hydric soils as saturated conditions are the
exception rather than the rule.

Functional guild — Categorization of plant communities into functional groups that can be defined by mea-
surable traits.

Gardians — “Cowboys” who ride horses through the wetlands of southern France’s Camargue.

Gator holes — Deep sloughs and solution holes that hold water during the dry season and that serve as wildlife
refuges; term mostly used in the Florida Everglades.

712 Glossary

Geogenous — Peatland subject to external flows.

Gleying — Development of black, gray, or sometimes greenish or blue-gray color in soils when flooded.

Glycolysis — See Fermentation.

Greenhouse gases (GHG) — ^Atmospheric gases that adsorb radiant energy at various wavelengths. The term
is mostly used to refer to the gases COj, CH4, and NjO that are products of human activity and that could
lead to atmospheric warming.

Guild — group of functionally similar species in a community.

HAB — Harmful algal bloom.

Halophiles — “Salt-loving” organisms .

Halophytes — Salt-tolerant plants.

Hammock — Slightly raised tree islands, such as tree island freshwater hammocks or mangrove islands in the
Florida Everglades.

Hatch-Slack-Kortschak pathway — Biochemical pathway of photosynthesis for C4 plants.

Hemists — Mucky peat or peaty muck; conditions between saprist and flbrist soil.

HGM — See Hydrogeomorphic classification.

High marsh — Upper zone of a salt marsh that is flooded irregularly and generally is located between mean
high water and extreme high water. Called inland salt marsh in Gulf of Mexico coastline.

Histosols — Organic soils that have organic soil material in more than half of the upper 80 cm, or that are of any
thickness if they overlie rock or fragmental materials that have interstices filled with organic soil material.

HLR — See Hydraulic loading rate.

HSI — Habitat suitability index, a semi-quantitative measure of habitat value of an ecosystem for specific species.

Hydrarch succession — Development of a terrestrial forested climax community from a shallow lake with wet-
land as an intermediate sere. In this view, lakes gradually fill in as organic material from dying plants
accumulates and minerals are carried in from upslope.

Hydraulic conductivity — See Permeability.

Hydraulic loading rate (HLR) — ^Amount of water added to a wetland, generally described as the depth of
water (volume of flooding per wetland area) per unit time; generally used for treatment wetlands.

Hydric soils — Soils that formed under conditions of saturation, flooding, or ponding long enough during the
growing season to develop anaerobic conditions in the upper part.

Hydrochory — Seed dispersal by water.

Hydrodynamics — ^An expression of the fluvial energy that drives a system.

Hydrogeomorphic classification (HGM) — Wetland classification system based on type and direction of
hydrologic conditions, local geomorphology and climate.

Hydrogeomorphology — Combination of climate, basin geomorphology, and hydrology that collectively influ-
ences a wetland’s fimction.

Hydroperiod — The seasonal pattern of the water level of a wetland. This approximates the hydrologic signature
of each wetland type.

Hydrophyte — Plant adapted to the wet conditions.

Hydrophytic vegetation — Plant community dominated by hydrophytes.

Hypoxia — Waters with dissolved oxygen less than 2mg/L.

Interception — Precipitation that is retained in the overlying vegetation canopy.

Intermittendy exposed — Refers to nontidal wetlands that are flooded throughout the year, except during
periods of extreme drought.

Intermittendy flooded — Refers to nontidal wetlands that are usually exposed, with surface water present for
variable periods without detectable seasonal patterns.

Intertidal — Part of coastal wetland flooded periodically with tidal water.

Intrariparian continuum — The structure and function of riparian communities along a river system.

Irregularly exposed — Refers to coastal wetlands with surface exposed by tides less often than daily.

Glossary 713

Irregularly flooded — Refers to coastal wetlands with surface flooded by tides less often than daily.

Isolated wetland — Legal term used in the United States to define wetlands that do not have an obvious
surface-water connection to a navigable stream or river {see also Significant nexus).

Jurisdictional wetland — Term used in the United States to refer to wetlands that fall under the jurisdiction of
federal laws for the purpose of permit issuance or other legal matters.

Kahikatea — Refers to both the tree (Dacrycarpiis dacrydiodes) and the forested wetlands found throughout
New Zealand. Referred to as “white pine” forests by locals.

Karst — topography formed over limestone, dolomite, or gypsum.

Krefeld system or Max-Planck- Institute process — Gravel bed macrophyte subsurface flow treatment wet-
lands.

Lacustrine — Pertaining to lakes or lake shores.

Lagoon — Term frequently used in Europe to denote deepwater enclosed or partially opened aquatic system,
especially in coastal delta regions.

Lentic — Related to slow-moving or standing water systems; usually refers to lake (lacustrine) and stagnant
swamp systems.

Lenticels — Small pores found on mangrove tree prop roots and pneumatophores above low tide and presumed
to be sites of oxygen influx for anaerobic roots survival.

Limnogenous peatland — Geogenous peatland that develops along a slow-flowing stream or a lake.

Littoral — Zone between high and low tide in coastal waters or the shoreline of a freshwater lake.

Loading rate — The amount of a material (e.g., a chemical) applied to a wedand, measured either per unit area
(e.g., g m“^ yr ') or volumetrically (e.g., g m“^ yr^')-

Lotic — Pertaining to running water (i.e., rivers and streams).

Low marsh — Intertidal or lower marsh in salt marsh that is located in the intertidal zone and is flooded daily.
Called streamside salt marsh in coastal Gulf of Mexico.

Mangal — Same as mangrove.

Mangrove — Subtropical and tropical coastal ecosystem dominated by halophytic trees, shrubs, and other plants
growing in brackish to saline tidal waters. The word “mangrove” also refers to the dozens of tree and shrub
species that dominate mangrove wetiands.

Marginal value — The value of an additional increment of a commodity in a free market.

Marsh — frequently or continually inundated wetland characterized by emergent herbaceous vegetation
adapted to saturated soil conditions. In European terminology, a marsh has a mineral soil substrate and
does not accumulate peat. See also Tidal freshwater marsh. Salt marsh.

Mesotrophie peatlands — ^Also called transition or poor fens. Peatlands intermediate between minerotrophic
and ombrotrophic.

Methane emissions — ^Amount of methane released from a landscape as net result of methanogenesis minus
methane oxidation.

Methane oxidation — Conversion by methane to methanol, formaldehyde, and carbon dioxide by obligate
methanotrophic bacteria.

Methanogenesis — Carbon process under extremely reduced conditions when certain bacteria (methanogens)
use COj or low-molecular-weight organic compounds as electron acceptors for the production of gaseous
methane (CH4).

Methanogens — Bacteria that carry out methanogenesis.

Methanotrophs — ^Aerobic bacteria that oxidize methane.

Millennium Ecosystem Assessment — International study published in 2005 that focused on the changes that
humans have caused to ecosystems and how those are affecting the services that they provide to humans.

Mineral soil — Soil that has less than 20 to 35 percent organic matter.

Minerotrophic peatlands — ^AIso called rheotrophic peatlands or rich fens; peatlands that receive water that has
passed through mineral soil.

714 Glossary

Mire — Synonymous with any peat-accumulating wetland (European definition); from the Norse word “myrr.”
The Danish and Swedish word for peatland is now “mose.”

Mitigate — To lessen or compensate for an impact. Used here in the context of mitigating wetland loss by
restoring or creating wetlands.

Mitigation bank — wetland area that has been restored and protected to provide compensation for impacts
to wetlands.

Mitigation ratio — The ratio of restored or created wetland to wetland lost to development.

Mitigation wetland — See Replacement wetland.

Moor — Synonymous with peatland (European definition). A highmoor is a raised bog; a lowmoor is a peatland
in a basin or depression that is not elevated above its perimeter. The primitive sense of the Old Norse root
is “dead” or barren land.

Mottles (or redox concentrations) — Orange/reddish-brown (because of iron oxides) or dark reddish-brown/
black (because of manganese oxides) accumulations in hydric soils throughout an otherwise gray (gleyed)
soil matrix. Motdes suggest intermittently exposed soils and are relatively insoluble, enabling them to
remain in soil long after it has been drained.

Muck — Sapric organic soil material with virtually all of the organic matter decomposed, not allowing for the
identification of plant forms. Bulk density generally greater than 0.2g/cm^ (more than peat).

Mimsell soil color chart — Book of standard color chips for determining soil color value and chroma. Used to
identify hydric soils.

Muskeg — Large expanse of peatlands or bogs; particularly used in Canada and Alaska.

NAD — Nicotinamide adenine dinucleotide, an enzyme that accumulates in anaerobic conditions.

NADP — NAD phosphate.

Nexus — Legal term emphasized by the U.S. Supreme Court to describe connections between wetlands and nav-
igable waterways that are regulated. Wetlands must significantly affect the chemical, physical, and biological
integrity of waters understood as “navigable” to have a significant nexus to those waters.

Nitrification — ^Ammonium nitrogen oxidized by microbes to nitrite nitrogen and nitrate nitrogen.

Nernst equation — Equation based on a hydrogen scale showing how redox potential is related to the concen-
trations of oxidants and reductants in a redox reaction.

Nitrogen fixation — Process in the nitrogen cycle whereby Nj gas is converted to organic nitrogen through
the activity of certain organisms in the presence of the enzyme nitrogenase.

No net loss — Wetland policy in the United States that began in the late 1980s and means that if wetlands are
lost they must be replaced so that there is no “net loss” of wetlands overall.

Nutrient budget — Mass balance of a nutrient in an ecosystem.

Nutrient spiraling — The process whereby resources (organic carbon, nutrients, etc.) are temporarily stored,
then released as they “spiral” downstream from organic to inorganic form and back again.

Obligate — Requiring a specific environment to grow, as in adapted only a wet environment. In the context of
wetlands, obligate generally refers to plants requiring saturated soils.

Oligotrophic — Nutrient poor; generally used in lake classification, but is also applicable to peatlands.

Oligotrophication — Often the process of peadand development whereby a peatland eventually elevates itself
above the surrounding landscape and goes from eutrophic (nutrient rich) to oligotrophic (nutrient poor).

Ombrogenous — Peatland with inflow from precipitation only; also called ombrotrophic.

Ombrotrophic — Literally rain fed, referring to wetlands that depend on precipitation as the sole source of
water.

Opportunity cost — The net worth of a non-free market resource in its best alternative use; that is, the net
benefit of the area in its best alternative use that has to be forgone in order to keep it in its natural state.

Organic soil — Soil that has more than 12 to 18 percent organic carbon, depending on clay content {seeFig. 06).

Osmoconformers — Marine animals in which the internal cell environment follows closely the osmotic concen-
tration of the external medium.

Glossary 715

Osmoregulators — Marine animals that control their internal cell environment despite a different osmotic con-
centration of the external medium.

Outwelling — Function of coastal wetlands as “primary production pumps” that feed large areas of adjacent
waters with organic material and nutrients; analogous to upwelling of deep ocean water, which supplies
nutrients to some coastal waters from deep water.

Overland flow — Nonchannelized sheet flow that usually occurs during and immediately following rainfall or a
spring thaw, or as tides rise in coastal wetlands.

Oxbow — ^Abandoned river channel, often developing into a swamp or marsh, on a river floodplain.

Oxidation — Chemical process of giving up an electron (e.g., Fe^’*' -» Fe^'*' + e“). Special cases involve uptake
of oxygen or removal of hydrogen (e.g., H^S ^ -F2H’'').

Oxidized pore linings — See Oxidized rhizosphere.

Oxidized rhizosphere (also called oxidized pore linings) — Thin traces of oxidized soils through an other-
wise dark matrix indicating where roots of hydrophytes were once found.

Paalsa peatlands — Peatlands found in the southern limit of the tundra biome; large plateaus of peat (20 to
100 m in breadth and length and 3 m high) generally underlain by frozen peat and silt.

Pakihi — Peatland in southwestern New Zealand dominated by sedges, rushes, ferns, and scattered shrubs. Most
pakihi form on terraces or plains of glacial or fluvial outwash origin and are acid and exceedingly infertile.

Palmer Drought Severity Index (PDSI) — relative measure of climatic “wetness.” Used primarily to esti-
mate the severity of droughts.

Paludifieation — The blanketing of terrestrial ecosystems by overgrowth of bog vegetation. See also Blanket
bog.

Palustrine — Nontidal wetlands.

Panne — Bare, exposed, or water-filled depression in a salt marsh.

Patterned fens — Aapa peatlands.

Peat — Fibric organic soil material with virtually all of the organic matter allowing for the identification of plant
forms. Bulk density generally less than O.lg/cm^ (less than muck).

Peatland — ^A generic term of any wetland that accumulates partially decayed plant matter (peat).

Penman equation — Empirical equation for estimating evapotranspiration using an energy budget approach.

Perched wedand — Wetland that holds water well above the groundwater table.

Permanently flooded — Refers to nontidal wetlands that are flooded throughout the year in all years.

Permeability — The capacity of soil to conduct water flow. Also known as hydraulic conductivity. See also Darcy’s
law.

Petagram (Pg) — 10^® grams.

Phreatophytes — Plants that obtain their water from phreatic sources (i.e., groundwater or the capillary fringe
of the groundwater table).

Physiognomy — The appearance or life form of vegetation.

Piezometers — Groundwater wells that are only partially screened and thus measure the piezometric head of an
isolated part of the groundwater.

Playa — ^An arid- to semiarid-region wetland that has distinct wet and dry seasons. Term used in the southwest
United States for shallow depressional recharge wetlands occurring in the Great Plains region of North
America that are formed through a combination of wind, wave, and dissolution processes.

Pnemnatophores — “Air roots” that protrude out of the mud from the main roots of wetland plants such as
black mangroves ( Avicennia) and cypress ( Taxodium distichum) and are thought to be organs for transport
of oxygen and other gases to and from the roots of the plant. Called “knees” for cypress.

Pocosin — Peat-accumulating, nonriparian freshwater wetland, generally dominated by evergreen shrubs
and trees and found on the southeastern Coastal Plain of the United States. The term comes from the
Algonquin for “swamp on a hill.”

Porosity — Total pore space in soil, generally expressed as a percentage.

716 Glossary

Pothole — Shallow marshlike pond, particularly as found in the Dakotas and central Canadian provinces; the
so-called prairie pothole region.

Prairie pothole — See Pothole.

Produeer surplus or economie rent — The area over a good’s supply curve bounded by price.

Prop roots — ^Above-ground arched roots that aid in support of some wetland trees such as the mangrove
Rhizophora.

Pulse stability eoncept — Concept that pulses can be both a subsidy and a stress to an ecosystem, depending
on their strength, with subsidies occurring with moderate pulses, while both weak and excessive pulses can
result in stress responses.

Quaking bog — Schwin^moor in German. Bog in which the peat layer and plant cover is only partially attached
in the basin bottom or is floating like a raft.

Quiekflow — Direct runoff component of streamflow during a storm that causes an immediate increase in
streamflow.

Raised bogs — Peat deposits that fill entire basins, are raised above groundwater levels, and receive their major
inputs of nutrients from precipitation. See Ombrogenous and Ombrotrophic.

Bamsar Convention — International treaty originally started in Bamsar, Iran, in the early 1970s to protect
wetland habitat around the world, especially for migratory waterfowl.

Baupo swamp — Cattail ( Typha) marsh in New Zealand.

Recharge wetland — ^Wetland that has surfacewater (or groundwater) level higher hydrologically than the sur-
rounding water table, leading to an outflow of groundwater.

Recurrence interval — The average interval between the recurrence of floods at a given or greater magnitude.

Redox concentrations — Bodies of accumulated iron and manganese oxides in wetland soils such as nodules
and concretions, masses (formerly called “reddish mottles”), and pore linings (formerly called “oxidized
rhizosphores”).

Redox depletions — Bodies of low chroma (2 or less) where the natural (gray or black) color of the parent sand,
silt, or clay results when soluble forms of iron, manganese, or clay are leached out of the soil. Generally
have Munsell color values of 4 or greater. See also Clay depletions.

Redoximorphic features — Features formed by the reduction, translocation, and/or oxidation of iron and man-
ganese oxides; used to identify hydric soils. Formerly called mottles and low-chroma colors.

Redox potential — Reduction-oxidation potential, a measure of the electron pressure (or availability) in a solu-
tion or measure of the tendency of soil solution to oxidize or reduce substances. Low redox potential
indicates reduced conditions; high redox potential indicates oxidized conditions.

Reduced matrix — Soil that has low chroma and high value but whose color changes in hue or chroma when
exposed to air.

Reduction — Chemical process of gaining an electron (e.g., Fe^'*' -t e“ -> Fe^'*'). Special cases involve releasing
oxygen or gaining hydrogen (hydrogenation) (e.g., S^“ +2U''' -» HjS).

Reedmace swamp — Cattail ( Typha) marsh in the United Kingdom.

Reedswamp — Marsh dominated by Phra^mites (common reed); term used particularly in Europe.

Reference wetland — Natural wetland used as a reference or control site to judge the condition of another
created, restored, or impacted wetland.

Regularly flooded — Refers to coastal wetlands with surface flooded and exposed by tides at least once
daily.

Regulators (or avoiders) — In reference to biological adaptations to stress, organisms that actively avoid stress
or modify it to minimize its effects.

Rehabilitation — Less than full restoration of an ecosystem to its predisturbance condition.

Renewal rate — See Turnover rate.

Replacement value — The sum of the cheapest way of replacing all the various services performed by a natural
ecosystem area.

Glossary 717

Replacement wetland — wetland constructed to replace the functions lost by human development, usually
in the same or an adjacent watershed.

Residence time — See Retention time.

Resource spiraling — See Nutrient spiraling.

Restoration — To return a site to an approximation of it condition before alteration. Se« a/to Wetland restoration.

Retention rate — The amount of a material retained in a wetland per unit time and area; usually refers to material
more or less removed from water flowing over or through a wetland, as contrasted to detention, which is
transitory.

Retention time — measure of the average time that water remains in the wetland. Nominal residence time or
retention time refers to the theoretical time that water stays in a wetland as calculated from the flowthrough
and the water volume in the wetland. See also Detention time.

Rheotrophic peatlands — See Minerotrophic peatlands.

Riparian — Pertaining to the bank of a body of flowing water; the land adjacent to a river or stream that is, at
least periodically, influenced by flooding.

Riparian ecosystem — Ecosystem with a high water table because of proximity to an aquatic ecosystem, usually
a stream or river. Also called bottomland hardwood forest, floodplain forest, bosque, riparian buffer, and
streamside vegetation strip.

River continuum concept (RCC) — Theory to describe the longitudinal patterns of biota found in streams
and rivers.

Root zone method ( Wurzelraumentsorgung) — Subsurface flow wetland basins, almost always found in
Europe, and generally planted with Phragmites australis.

Runoff — Nonchannelized surfacewater flow.

Salt exclusion — ^A salinity adaptation by some wetland plants by which plants prevent salt from entering the
plant at the roots.

Salt marsh — ^A halophytic grassland on alluvial sediments bordering saline water bodies where water level fluc-
tuates either tidally or nontidally.

Salt seeretion — ^A salinity adaptation by which some wetland plants excrete salt from specialized organs in the
leaves.

Saprists — See Muck.

Saturated soils — Refers to nontidal wetlands where the soil or substrate is saturated for extended periods in
the growing season, but standing water is rarely present.

Sclerophylly — Refers to the thickening of the plant epidermis.

Seasonally flooded — Refers to nontidal wetlands that are flooded for extended periods in the growing season,
but with no surface water by the end of the growing season.

Seca — Dry period in Pantanal region of South America from September through November when the wetland
reverts to vegetation typical of dry savannas. See also Cheia, Enchente, and Vazante.

Secondary treatment — Treatment of wastewater to remove organic material.

Sedge meadow — ^Very shallow wetland dominated by several species of sedges (e.g., Carex, Scirpus, Cyperus).

Seed bank — Seeds stored in soils, often for many years. In wetlands changing hydroperiod, as in wetland
restoration or wedand drainage, can often lead to germination.

Self-design — The application of self-organization in the design of ecosystems. The process of ecosystem devel-
opment whereby the continual or periodic introduction of species propagules (plants, animals, microbes)
by humans or nature and their subsequent survival (or nonsurvival) provide the essence of the successional
and functional development of an ecosystem.

Semipermanently flooded — Refers to nontidal wetlands that are flooded in the growing season in most years.

Sequestration — The permanent retention of a chemical or nutrient in an ecosystem. Often used to describe
the permanent burial of carbon in wedand soils as carbon sequestration.

718 Glossary

Serial discontinuity concept — Describes the effects that floodplains, dams, and the transverse dimension in
general has on the functioning of a river system.

Shrub-scrub — Wetlands dominated by woody, low-stature vegetation such as freshwater buttonwood ( Cepha-
lanthus) or saltwater dwarf mangrove (Rhizophora) swamps.

Significant nexus — Legal term used in the United States to describe the connection of a wetland to an adja-
cent navigable water. The wetland should, by itself or in combination with other lands, significantly affect
the chemical, physical, and biological integrity of the adjacent navigable water {see also Isolated wetland;
Nexus).

Sink — Term used in the context of wetland nutrient budgets to define a wetland that imports more of a certain
nutrient than it exports.

Slough — ^An elongated swamp or shallow lake system, often adjacent to a river or stream. A slowly flowing
shallow swamp or marsh in the southeastern United States (e.g., cypress slough). From the Old English
word “sloh” meaning a watercourse running in a hollow. See also Cypress strand.

Soligenous peatland — Geogenous peatland that develops with regional interflow and surface runoff.

Source — Term used in the context of wetland nutrient budgets to define a wetland that exports more of a
certain nutrient than it imports.

Spit — A neck of land along a coastline behind which coastal wetlands sometimes develop.

SRP — Soluble reactive phosphorus, similar to orthophosphate; a measure of biologically available phosphorus.

Stem hypertrophy — Noticeable swelling of lower stem of vascular plant, usually caused by water or saturated
soils. Includes tree buttresses and fluted trunks.

Stemflow — Precipitation that passes down the stems of vegetation. Used generally in connection with forests
and forested wetlands.

Streamflow — Channelized surfacewater flow.

Stream order — Numerical system that classifies stream and river segments by size according to the order of its
tributaries.

String bogs — Aapa peatlands.

Strings — Aapa peatlands.

Subsidence — Sinking of ground level, caused by natural and artificial setding of sediments over time.

Subsurface-flow constructed wetlands — Constructed wetlands through which water flows beneath the sur-
face rather than over the surface. See also Root zone method.

Subtidal — Coastal wetland permanently flooded with tidal water.

Supply curve — Economist’s estimate of producer benefits.

Surface-flow constructed wetlands — Constructed wetlands that mimic many natural wedands with flow on
surface rather than below the surface.

Swamp — ^Wedand dominated by trees or shrubs (U.S. definition). In Europe, forested fens and wetlands dom-
inated by reed grass {Phragmites) are also called swamps {see Reedswamp).

Swampbuster — Provision of the U.S. Food Security Act that encourages farmers not to drain wetlands and
thereby lose their farm subsidies.

Swamp gas (or marsh gas) — Methane.

Taking — The legal denial of an individual’s right to use all or part of the area or structure (trees, wildlife, etc.)
of his or her property.

Telmatology — ^A term originally coined to mean “bog science.” From the Greek word “telma” for bog.

Temporarily flooded — Refers to nontidal wetlands that are flooded for brief periods in the growing season,
but otherwise the water table is well below the surface.

Teragram (Tg) — 10'^ grams; a billion kilograms.

Terrestrialization — Generally in reference to succession of peatlands, the infilling of shallow lakes until they
become, in appearance, a peat basin supporting terrestrial vegetation.

Glossary 719

Tertiary treatment — ^Advanced treatment of wastewater after secondary treatment to remove inorganic nutri-
ents and other trace materials. Wetlands are often used for this purpose.

Thornthwaite equation — Empirical equation for estimating potential evapotranspiration as a function of air
temperature.

Throughfall — Precipitation that passes through vegetation cover to the water or substrate below. Used partic-
ularly in forests and forested wetlands.

Tidal creeks — Small streams that serve as important conduits for material and energy transfer between salt
marshes or mangrove swamps and adjacent coastal bodies of water.

Tidal freshwater marsh — Marsh along rivers and estuaries close enough to the coastline to experience signif-
icant tides by nonsaline water. Vegetation is often similar to non tidal freshwater marshes.

Tolerators (also called resisters) — In reference to biological adaptations to stress, organisms that have func-
tional modifications that enable them to survive and often to ftmction efficiently in the presence of stress.

Topogenous — Refers to peatland development when the peatland modifies the pattern of surface water flow.

Total suspended sohds — See TSS.

Transformer — Term used in the context of wetland nutrient budgets to define a wetland that imports and
exports the same amount of a certain nutrient but changes it from one form to another.

Translocation — Movement of nutrients between below-ground and above-ground portions of plants.

Treatment wetland — ^Wetland constructed to treat wastewater or polluted runoff. See also Constructed wet-
land.

TSS — Total suspended solids, a measure of the sediments in a unit volume of water.

Turlough — Term is specific for these types of wetlands found mostly in western Ireland. Areas seasonally
flooded by karst groundwater with sufficient frequency and duration to produce wetland characteristics.
They generally flood in winter and are dry in summer and fill and empty through underground passages.

Turnover rate — Ratio of throughput of water to average volume of water within a wetland. This is the inverse
of turnover time, residence time, or retention time of a wetland.

Turnover time — See Retention time.

Value — Something worthy, desirable, or useful to humanity; although the term is used often in ecology to
refer to processes (e.g., primary production) or ecological structures (e.g., trees) as they are “valuable” to
the way an ecosystem ftmctions, the term generally should be limited to an anthropocentric connotation.
Humans decide what is of “value” in an ecosystem.

Vazante — Period of declining water in Pantanal region of South America from June through August. See also
Cheia, Enchente, and Seca.

Vernal pool — Shallow, intermittently flooded wet meadow, generally typical of Mediterranean-type climate
with dry season for most of the summer and fall. Term is now used to indicate wetlands temporarily
flooded during the spring throughout the United States.

Viviparity — The production of young in a living state.

Viviparous seedlings — Seedlings of trees germinate while still attached to the tree canopy, as with the man-
grove genera Rhizophora. A specific case of viviparity.

Vleis — Seasonal wetland similar to a Dambo; term used in southern Africa.

Wad (pi. Wadden) — Unvegetated tidal flat originally referring to the northern Netherlands and northwestern
German coastline. Now used throughout the world for coastal areas.

Watertraeks — See Aapa peadands.

Wetland — See various wetland definitions in Chapter 2. Generally, wetlands have the presence of shallow water
or flooded soils for part of the growing season, have organisms adapted to this wet environment, and have
soil indicators of this flooding such as hydric soils.

Wetland ereation — The conversion of a persistent upland or shallow water area into a wetland by human
activity.

Wetland delineation — The demarcation of wetland boundaries for legal purposes. See Jurisdictional wetlands.

720 Glossary

Wetland Reserve Program (WRP) — a U.S. government policy for farmers to receive benefits to set aside wet-
lands on their property for conservation.

Wetlanders — People who live in proximity to wetlands and whose culture is linked to the wetlands.

Wedand restoration — The return of a wetland from a condition disturbed or altered by human activity to a
previously existing condition.

Wedands of international importance — Wetlands designated by the Ramsar Convention as important inter-
national wetlands because they contain rare wetland types, support biological diversity, waterfowl, and
fish.

Wet meadow — Grassland with waterlogged soil near the surface but without standing water for most of the
year.

Wet prairie — Similar to a marsh, but with water levels usually intermediate between a marsh and a wet meadow.

Willingness-to-pay, or net willingness-to-pay — hypothetical market that establishes the amount society
would be willing to pay to produce and/or use a good beyond that which it actually does pay.

Reference

Odum, H. T. 1996. Environmental Accounting: Emer^y and Environmental Decision Making. John Wiley &
Sons, New York. 370 pp.

Index

Aapa pcatlands, 420-421
Acer rubrum (red maple), 59, 376
in red maple swamps, 388
root adaptations, 220
root systems, 395
swamp productivity/hydrology
relationship, 147
Acidification, 432

Actophilornis africanus (African jacana),

Adaptations;

of freshwater swamp vegetation, 391-395
of mangrove vegetation, 322-326
morphological, 216-220
mutualism and commensalism, 226-227
of peatland vegetation, 431-433
physiological, 220-226
of vascular plants, 216-227
whole plant strategies, 226
Adenosine triphosphate (ATP), 226
ADH (alcohol dehydrogenase), 224
Adventitious roots, 218, 219, 227, 395
Acrcnchyma, 216-218
Aerial habitat, 274-275
Aerobic bacteria, 281
Aerobic respiration, 195
Aesthetics of ecosystem services, 545
African jacana [Actophilornis africanus)^ 85
Alxican reef heron [Egretta gularis)^ 88
Alfican wetlands, 56, 83-89
Congolian swamp forests, 86
East Africa tropical marshes, 86-88
Okavango Delta, 84-86
West Africa mangrove swamps, 87-89
Agkistrodon piscivorus (cottonmouth
moccasin), 396-397

Agricultural Conservation Program (U.S.

Department of Agriculture), 481
Agricultural land restoration, 598
Agricultural stormwater treatment wetlands,
659-662

Agricultural wastewater wetlands,

662-663

Agriculture:

in Nebraska sandhills, 62
wetland management by objective for,
497-499
Alaska;

black spruce pcatlands, 430-431
exclusion of wetlands from Food Security
Act definition, 40, 41
wetlands acreage, 49
Alder (Alnus)^ 82
gas transport in, 223
mutualism, 227
Algae:

and decomposition process, 282-283
in freshwater tidal marshes, 290
and nitrogen fixation, 187, 188
in salt marsh food web, 277-279
Alligator mississippiensis (American
alligator), 114, 396, 537-538
Allogenic succession, 215, 228, 230
autogenic succession vs., 227-235
community concept and continuum idea,
232-234

Lake Michigan dunes, 230-232
linear directed change, 234
seed banks, 234-235
Alluvial cypress swamp, 380
Alluvial river swamp, 379, 381
Alnus (alder), 82, 223, 227

Alnus barbata^ 82

Alnus glutinosa^ 82, 223

Alopochen aegyptiaca (Egyptian goose),

Alteration of wetlands, 482-492
by drainage, dredging, and filling,
482-486

for highway construction, 487-488
hydrologic modifications, 486-487
by mineral and water extraction,

490-491

for peat mining, 488-489
by water pollution, 491
Amazon basin, 77-78
Amazon River, 78, 119
American alligator [Alligator

mississippiensis)^ 114, 396, 537-538
American bulrush [Scirpus americanus)^ 67,
68

American eel [Anguilla rostrata)^ 292
Ammonia, 198

Ammonia transformations, 183-184
Ammonia volatilization, 183, 208, 209
Ammonification, 181. 5^^ /t/jo Nitrogen
mineralization

Ammonium nitrogen, 153, 183-184
Amphibians;

freshwater marshes, 355-356
freshwater swamps, 396-397
peatlands, 434

in tidal freshwater wetlands, 294-295
Anadromous species, 292
Anaerobic bacteria, 281
Anaerobic processes, 195-196
Anaerobic respiration, 224-226
Anaerobiosis, 279-280

722 Index

Anammox (anaerobic ammoiiium
oxidation), 188
Anas (dabbling ducks), 531
Anas erythrorhynchos (red-billed duck),

Anas platyrhynchos (mallard), 532
Anas undulata (yellow-billed duck), 88
Anguilla rostrata (American eel), 292
Anishinaabe, 7
Anoxia, 215, 220-221
Anoxygenic photosynthesis, 195
Aquaculture, 497-499
Aquatic civilizations, 5, 479
Aquatic habitat, 276
Aquic condition, 169-170
Aquifer recharge, 544
Archaea^ 197
Arctic salt marshes, 262
Ardea cincrea (gray heron), 533
Ardea cocol (white-necked heron), 533
Ardea herodias (great blue heron), 533
Ardea melanocephala (black-headed heron),
533

Ardea novaehollandiae (white-faced heron),
533

Arid riparian ecosystems:
geographic extent, 377
geomorphology and hydrology, 384
vegetation, 390
Arid riparian forests, 390
Ai'tificial management, 493, 494
Artificial wetland, 49
Asian wetlands, 56, 93-100
China, 95-100

Indian freshwater marshes, 93, 94
Issyk Kill, 95

southern Asia river deltas, 93-95
Taiwan urban wetland park, 99, 100
western Siberian lowlands, 93
Assiniilatory nitrate reduction, 184
Assiniilatory sulfate reduction, 194
Atchalalaya Basin, 65
Atchafalaya River, 65
Atchalalaya Swamp (Louisiana), 6
Athabasca River, 70
Atlantic coastal plain:
swamp rivers of, 60-61
tidal salt water marshes, 264
Atlantic white cedar {Chamaecyparis
thyoides)^ 59, 375-376
ATP (adenosine triphosphate), 226
Australasia:

loss of wetlands in, 50
salt marshes, 265
tidal salt marshes of, 265
wetlands loss, 50
Australian wetlands, 56, 89-91
eastern billabongs, 89-90
losses of, 50

tidal salt water marshes, 265
western wetlands, 90-91
Autochthonous inputs, 344
Autogenic succession, 215, 227-228
allogenic succession vs., 227-235
community concept and continuum idea,
232-234

Lake Michigan dunes, 230-232
linear directed change, 234
seed banks, 234-235
Avicennia (black mangrove), 314, 325
adaptations, 324
in basin mangroves, 317
plant zonation, 321
and pneumatophores, 220, 325
rhizosphere oxygenation, 224
salinity control, 323
speciation, 312

Avicennia ^erminans^ 314, 325

Bacillariophytes, 290
Backswamp, 382
Bacteria:

and anoxygenic photosynthesis, 195
as decomposers, 281
and niethanogenesis, 197
and sulfate reduction, 194
and sulfur cycle, 201
Bald cypress {Taxodium distichum)^ 220,
374, 375, 387
Bangladesh, 94-95
Banldull discharge, 133
Barataria Bay estuary, Louisiana, 66
Barlctt, Charles, 63
Basin mangroves, 317, 318
Basin morphology, 670-671
Basin wetland, 318, 320
Bay of Bengal Delta, 94-95
Bay of Fundy:
salt marshes, 264
tidal salt water marshes, 264
Beaver {Castor canadensis)^ 114, 295
pelts from, 530-531
as threat to reconstructed wedand, 682
Bengal tiger {Panthera ti^ris tifjris)^ 95
Benthic habitat, 276
Beowulf^ 9

Berezinski Bog, Byelorussia, 82-83
Big Cypress Swamp, Florida, 57, 119,

120

Billabongs, 33, 34, 89-90, 381, 382
Bi-Ob region (Russia), 93
Biogeochemical cycling, 179
Biogeochemistry, 179-212
ammonia transformations and
nitrification, 183-184
anammox, 188
carbon cycle, 195-202
carbon-sulfur interactions, 201-202

dissimilatory nitrate reduction to
ammonia, 188

ebullition and gaseous transport in plants,
200-201

fermentation, 196-197
freshwater marshes, 344-348
freshwater swamps, 384-385
and hypoxia, 189-192
inflows and outflows, 180-181
intrasystem cycling, 180
iron and manganese transformations, 192
methane emissions, 198-200
methane oxidation, 197-198
niethanogenesis, 197
nitrate transformations and
denitrification, 184-187
nitrogen cycle, 181, 183-192
nitrogen fixation, 187-188
nitrogen mineralization, 181
nutrient budgets, 208-212
oceans and estuaries, 205
peatlands, 424-427
phosphorus cycle, 202-204
sinks, sources, and transformers of
chemicals, 181, 182
streams, rivers, and groundwater,
206-207

sulfur cycle, 193-195
water quality, 204-207
Biotic control of hydrology, 113-115
Birds:

freshwater marshes, 357-358
peatlands, 434-435
as provisioning ecosystem service,
531-534

in tidal freshwater wetlands, 293-294
Blackbird (Icteridae), 294, 358
Black-headed heron {Ardea melanocephala)^
533

Black mangrove {Avicennia):
adaptations, 324
in basin mangroves, 317
plant zonation, 321
and pneumatophores, 220, 325
rhizosphere oxygenation, 224
speciation, 312

Black spruce {Picea mariana)^ 430
Black spruce-feathermoss forest, 428
Black spruce peatlands, 430-431
Blanket bogs, 422
Blue crab {Callinectes sapidus)^ 284
Blue-green algae, 188, 290
BOD (biochemical oxygen demand), 650
Bog, 33, 34,413

Bog energy flow estimates, 440-442
Bog-iron, 192
Bottomland, 34

Bottomland hardwood forest, 118, 374,

377

Index 723

Branta canadensis {Canada goose), 114,
532, 682

Building materials, 10-14
Bulk density, 163
Burmese python {Python molurus
bivittatus)^ 601-602
Buttonwood {Conocarpus erecta)^ 321
Buttresses, 218, 219, 394
Byrd, William, 111, 8

Cajuns, 5
California:

algal productivity on coast, 279
selenium contamination at Kesterson
National Wildlife Refuge, 401
wetland loss, 16, 66, 485
Callinectes sapidus (blue crab), 284
Calusa, 6

Camargue (France), 5, 6, 78, 79
Canada goose {Branta canadensis)^ 114,
532,682

Canadian Wetland Classification System,
423-424, 465,466
Canadian wetlands, 56
definitions of, 37-38
Great Lakes Wetlands/St. Lawrence
Lowlands, 66-68
Hudson-James Bay lowlands,

68-70

losses of, 50, 54-55
Peace-Athabasca Delta, 70-71
peatland restoration, 605-607
peatlands of Canada’s central and eastern
provinces, 68
prairie potholes, 61-62
Capital costs of treatment wetlands,
688-689

Cap Tourmente freshwater marsh complex
(Canada), 66-68

Carabell v. U.S. Army Corps of En_gineers^
519

Carbon budget, peatlands, 443-445, 568,
578-579

Carbon cycle, 195-202. See also Global
carbon cycle

carbon-sulfur interactions, 201-202
ebullition and gaseous transport in plants,
200-201

fermentation, 196-197
methane emissions, 198-200
methane oxidation, 197-198
methanogenesis, 197
Carbon export, 151-153
Carboniferous, 3
Carbon sequestration, 568-570
Carbon storage:

in mangroves, 330-332
in peatlands, 444, 567, 579
Carbon-SLilfrir interactions, 201-202

Carex (sedge);

in peat composition, 164
swamp productivity/hydrology
relationship, 147
Carnivorous plants, 432-433
Carolina pocosins, 59, 431
Caspian sea, 82

Castor canadensis (beaver), 114, 295
pelts from, 530-531
as threat to reconstructed wetland, 682
Catadromous species, 292
Cat clays, 319
Cation exchange, 426
Cation exchange capacity, 165, 166
Cattail {Typha)^ 73, 74

adaptations to waterlogging/flooding,

222

in centrifugal organization model, 239
in colonized wetlands, 635
in Everglades, 601, 602
and flow-through, 146
freshwater marsh zonation, 349
and nutrient enrichment, 659-660
nutrient limitations, 365
photosynthetic efficiency, 359-360
in restored wetlands, 635
and upland soils, 627-630
and wetland primary succession, 246-249
Cedarburg Bog (Wisconsin), 421
Cedar string bog and fen complex, 428
Central American wetlands, 72-74
Centrifrigal organization concept
(succession), 238-240
Chamaecyparis thyoldes (white cedar), 59,
375-376, 394, 395
Channelized streamflow, 130-132
Cheia, 76
Chemical loadings:

design graphs, 671-672
empirical models, 675-676
retention rates, 673-674
in treatment wedands design, 671-676
Chemosynthetic bacteria, 192
Chen (snow goose), 114, 682
China wetlands, 95-100
loss of, 49, 50
northeastern wetlands, 97
Qinghai-Tibetan Plateau, 97, 98
river deltas, 95, 96
urban wedand parks, 98-99
Yangtze River wetlands, 95, 97
Chlorophytes, 290
Chongming Island, China, 95, 96
CICOLMA (Coastal Research Center La
Mancha, Veracruz, Mexico), 72
Circular 39 classification (U.S. Fish and
Wildlife Service), 457-459
Cladium (sedge), 147
Cladium jamaicense (sawgrass), 57, 602

Classification of wetlands, see Wetland
classification

Classification of Wetlands and Deepwater
Habitats, 460-464, 466
Clay:

depletions, 168, 170
and phosphorus cycle, 204
phosphorus sorption onto, 204
Clean Water Act, 39, 40, 507-509, 518
Clean Water Act amendments (1977), 39
Climate change, 563-584
carbon budgets for peadands, 578-579
carbon sequestration, 568-570
causes of, 565-567
climate change feedbacks, 577-578
coastal wetlands, 579-582
effects on wetlands, 579-584
inland wetlands, 582-584
methane emissions, 570-577
peat storage and global carbon budget,
568

wetlands in global carbon cycle,
567-579

Climate regulation, 543-545
aquifer recharge, 544
water quality, 544-545
Climate zones, 47

Clinton, Bill, and administration, 509
Clonal dominant species, 236
Coastal marshes, 630. See also Tidal salt
marshes

Coastal Plain (United States), 264
Coastal protection, 540-543
Coastal Research Center La Mancha

(CICOLMA) (Veracruz, Mexico), 72
Coastal restoration, 607-622
after Indian Ocean tsunami, 618
Delaware Bay salt marsh, 608-613
deltas, 619

mangrove swamps, 616-618
New York City region, 614-616
rivers, 619-620

River Skjern, Denmark, 620-622
salt marshes, 607-616
Coastal salt marsh, 165. See also Tidal salt
marshes

Coastal squeeze, 580
Coastal wetlands:

classification of, 459-460
effects of climate change on, 579-582
hydrology/productivity relationship,
150-151

Coastal wetland ecosystem, 456
Coastal Zone Management Program,

518

Coliforms, 663

Colorless sulfur bacteria (CSB), 194
Columbus, Christopher, 74
Commensalism, 226-227

724 Index

Common reed {Phra^mites australis)^ 82,
89,609,613

as invasive species, 353-354
and reed swamps, 342
in restored wetlands, 635
Community concept (succession), 232-234
Concentric domed bog, 420
Concretions, 170
Congolian swamp forests, 86
Conocarpus erecta (buttonwood), 321
Conservation of wetlands, 16-17
Conservation Reserve Program (CRP),

597

Constructed wetland:
classification of, 648
defined, 593
Consumers:

freshwater marshes, 354-358
freshwater swamps, 395-397
mangrove swamps, 326-327
peatlands, 433-435
tidal freshwater wetlands, 290-295
tidal salt marshes, 274-277
Consumption in tidal salt marshes, 281-283
Continuum idea (succession), 232-234
Convention of Wetlands, 39
Conversion factors, 709-712
Corkscrew Swamp, Greater Florida
Everglades, 198, 200
Cotton grass {Eriophorums)^ 432
Cottonmouth moccasin {A^kistrodon
piscivorus)^ 396-397
Cowles, H. C., 227, 230
Coypu {Myocastov coypus)^ 276-277
Crabs, 326
Cranberries, 9

Created wetland, 570, 571, 574. See also
Restoration and creation of wetlands
Cropped wetlands, 597
Crow Wing County, Minnesota, 7
CRP (Conservation Reserve Program), 597
CSB (colorless sulfur bacteria), 194
Cultural ecosystem services, 545
Cumbungi swamp, 33, 34
Cumulative loss, 32
Cyanobacteria, 188, 290
Cypress (Taxodium)^ 57

adaptations to waterlogging/flooding,
219

and buttresses, 394
longevity, 395

Cypress domes, 117, 378-380
Cypress strand, 379
Cypress swamps, 373
geomorphology, 377-381
hydrology/productivity relationship,

147

vegetation, 386-388
Cypress-tupelo swamps, 374-375

Dabbling ducks (Anas)^ 293, 357, 531
DAI (days after inundation), 198, 200
Dalton’s law, 137
Dambo, 34

Danube River delta, 82

Darcy’s law, 137

Davis, John Henry, 318-321

Days after inundation (DAI), 198, 200

Day water lily {Nymphaea nottchali var.

caerulea)^ 85
Decomposers, 281
Decomposition:

freshwater marshes, 361-362
of organic soils, 167
peatlands, 439

in tidal salt marshes, 281-283
Deepwater marsh, 33
Deepwater swamps, 598, 630
Defining wetlands, 27-42
common terms for, 33-36
by distinguishing features, 28-31
formal definitions, 36
legal definitions, 39-42
Deforestation, 77
Delaware Bay salt marsh, 608-613
Delineation, 28. Wetland

delineation guidelines (U.S.)

Deltas, 34, 619. See also specific deltas, e.^.:
Mississippi River Delta
Demand curve, 550
Denitrification, 184-187
climate regulation, 543
Gulf hypoxic zone, 189
nutrient budgets, 301
nutrient cycling, 153
peatlands, 442

as source of nitrous oxide, 567
treatment wetlands, 671
Designer wetland, 246, 633
Design graphs, 671-672
Des Plaines River Wetlands, 666-668
Desulfovibrio, 194
Detention time, 670
Detritivores, 326
Diatoms, 290

Die-off of salt marshes, 284-285
Dimethyl sulfide, 194
Dinitrogen (N^), 186
Discharge wetlands, 134-137
Dismal Swamp Land Company, 59
Dissimilatory nitrate reduction to ammonia
(DNRA), 188

Dissimilatory nitrogenous oxide reduction,
184

Dissolved organic phosphorus (DOP),

203

Distinguishing features of wetlands, 28-31
Diversion wetlands, see River diversion
wetlands

DMS (dimethyl sulfide), 194
DNRA (dissimilatory nitrate reduction to
ammonia), 188
Dominance types, 463
DOP (dissolved organic phosphorus),

203

Drainage:

alteration of wetlands by, 482-486
U.S. wetland drainage history, 480-482
Drawdown, 494
Dredging, 482-486
Drop roots, 325
Ducks, 293-294, 531-534
Duck potato, 226
Duck stamps, 16-17
Dwarf cypress swamps, 378, 379
Dwarf mangroves, 317-318

East African tropical marshes, 86-88
Eastern Asia:

salt marshes, 265

Eastern Australia billabongs, 89-90
Eastern North America:
salt marshes, 264
Eat-outs, 1 14

Ebro Delta, 14, 78-79, 278
Ebullitive flux (ebullition), 201
Ecological engineering, 647
Ecological principles, 638-639
Ecological success, 594
Ecological valuation of ecosystem services,
546-550

habitat evaluation procedures, 546
hydrogeomorphic analysis, 547-550
Economic evaluation of ecosystem services,
550-558

emergy analysis, 555-556
energy analysis, 553-555
opportunity costs, 551
replacement value, 551-553
valuing goods and services, 556-558
willingness-to-pay methods, 551
Economics of treatment wetlands, 688-691
additional benefits, 690
capital costs, 688-689
comparison of wetlands with

conventional technology, 690-692
operating and maintenance costs, 689
Ecosystem development, 240-251
ecosystem engineers, 249
landscape patterns, 250-251
pulse stability, 244-245
self-organization and self-design,
245-249

strategy of, 242-249
turnover rates and nutrient influxes,
242-244

wetland insularity, 244
Ecosystem engineers, 1 14, 249

Index 725

Ecosystem function:

bog energy flow estimates, 440-442
decomposition, 361-362, 439
decomposition and consumption in salt
marshes, 281-283
die-off of salt marshes, 284-285
energy flow, 402
food webs, 362
freshwater marshes, 358-366
freshwater swamps, 397-403
greenhouse gas emissions, 365-366
herbivory, 361-362
in mangrove swamps, 327-334
nutrient budgets, 362-364, 402, 403,
442-445

nutrient limitations, 364-365
organic export in salt marshes, 283-284
peat accumulation, 440
peatlands, 436-445

primary productivity, 277-281, 327-329,
358-361, 397-402, 436-438
tidal freshwater wetlands, 295-301
tidal salt marshes, 277-286
Ecosystem mass balance, 208
Ecosystem services, 527-559

animals harvested for pelts, 529-531
climate regulation, 543-545
cultural, 545

current paradigm of, 528-529
ecological valuation, 546-550
economic evaluation, 550-558
endangered and threatened species,
536-538

Faustian bargain in, 559
fish and shellfish, 534-535
flood mitigation, 539-540
peat harvesting, 536
problems and paradoxes of, 558-559
provisioning services, 529-538
quantifying, 545-559
regulation of, 539-543
storm abatement and coastal protection,
540-543

timber and vegetation harvest, 535-536
waterfowl and other birds, 531-534
Ecotones, 230
Ecotourism, 13-15

Egretta gularis (African reef heron), 88
Egyptian goose {Alopochen ae^yptiaca)^ 88
Eleocharis sphaceiata^ 199, 201
Elymus athericus^ 273
Emergency Wetlands Resource Act, 518
Emergy, 553

Emergy analysis, 555-556
Empirical models, 675-676
Enchente, 76

Endangered species, 434, 536-538
Endla Bog, Estonia, 82-83
Energy analysis, 553-555

Energy flow:

freshwater swamps, 402
in tidal freshwater wetlands, 297-298
Ensenada del Pabellon (Mexico), 71
Environmental sieve model, 236-238
Epiphytic algae, 282
Ericaceous plants, 415, 432, 436
Eriophorums (cotton grass), 432
Esox Indus (northern pike), 356
Estuaries:

effects on mangrove swamps, 332-334
water quality, 205
Estuarine system, 460, 461
Eucalyptus camaldulensis (river red gum),

Euprhates River, 89
European wetlands, 56, 78-83
loss of, 49, 50

Mediterranean Sea deltas, 78-79
northern Europe coastal marshes, mud
flats, and bays, 80-82
peatlands, 82-83
Rhine River delta, 79-80
southeastern Europe inland deltas, 82, 83
Eutrophic (term), 423
Eutrophic peatlands, 422
Evapotranspiration, 137-142
direct measurement of, 138-139
and ecosystem development, 228
effects of vegetation on, 141-142
empirical estimates of, 139-141
measuring, 154

Thornthwaite equation, 139-141
Excentric raised bogs, 417, 420
Executive Orders, 504, 506-508
Exotic species, 634-635
Exports, hydrologic, 208

Fen, 33, 34, 146,413
Fermentation:

in anaerobic respiration, 225, 226
in carbon cycle, 196-197
carbon-sulfur interactions, 201
Fern, 288
Ferrous iron, 192
Ferrous sulfide, 195
Fibrists, 167. See also VcdX
Fiddler crab ( Uca)^ 326
Filling of wetlands, 482-486
Filter feeders, 326
Finch (Fringillidae), 294
Fish:

freshwater marshes, 356-357
freshwater swamps, 396
as provisioning ecosystem service,
534-535
Flarks, 420,421
Floating marshes, 288
Floods, 132-134

Flood control:

hydrologic modifications for, 486-487
wetland management by objective for,
499

Flood duration, 115
Flood frequency, 115
Flood mitigation, 539-540
Flood peak, 129

Flood pulse concept (FPC), 404-406

Flordia Everglades, 6-8

Florida:

Big Cypress Swamp, 57, 119, 120
energy analyses of wetlands, 555-556
mangroves, 328
mangrove swamps, 313-334
rating wetlands in, 470-471
Florida Everglades, 9, 55-58, 492
restoration of, 598-603
treatment wetlands, 659-662
water quality restoration, 598-602
Florida panther {Puma concolor coryi)^ 58
Flowthrough succession, 419
Fluted trunk, 219
Flycatcher (Tyrannidae), 294
Folists, 167
Food, 9-10
Food Security Act, 39
Food Security Act definition of wetlands,
40-41

Food webs, 362
Forested wetlands:

introducing vegetation, 630
restoration of, 597-598
Formal definitions of wetlands, 36
FPC (flood pulse concept), 404-406
France, 5, 6, 78, 79
Frankia alni^ 227
Freedom Park, Florida, 656-659
Freshwater marshes, 341-366
biogeocheniistry, 344-348
consumers, 354-358
decomposition and herbivory, 361-362
ecosystem function, 358-366
food webs, 362

greenhouse gas emissions, 365-366
hydrology, 343-344
introducing vegetation, 630
invasive species, 353-354
nutrient budgets, 362-364
nutrient limitations, 364-365
primary productivity, 358-361
seed banks, 350-351
species diversity, 350-353
vegetation, 348-354
zonation, 348-349
Freshwater swamps, 373-403. See also
Riparian ecosystems
biogeocheniistry, 384-385
consumers, 395-397

726 Index

Freshwater swamps {^continued)

cypress swamps, 377-381, 386-388
cypress-tiipelo swamps, 374-375
ecosystem function, 397-403
energy flow, 402
fish, 396

geographic extent, 374-375
geomorphology and hydrology, 377-381
glaciated regions, 390-301
invertebrates, 395-396
nutrient budgets, 402, 403
primary productivity, 397-402
red maple swamps, 376, 381, 388-389
reptiles and amphibians, 396-397
tree adaptations, 391-395
vegetation, 385-395
white cedar swamps, 375-376, 381, 388
Freshwater tidal marshes, 259
Fringe mangroves, 316-318
Functional groups, 236
Functional guilds, 237
Fungi, 281

Gandau Nature Park {Taipei, Taiwan),
99-100

Ganges Delta, 94
Gardians, 78
Gaseous diffusion, 395
Gator holes, 114, 537
Geogenous peatlands, 423, 424
Geomorphology:
arid ecosystems, 384
cypress swamps, 377-381
freshwater swamps, 377-381
mesic ecosystems, 382-383
red maple swamps, 381
riparian ecosystems, 381-384
white cedar swamps, 381
Geukensia demissa (ribbed mussel), 282
Glaciated regions, freshwater swamp
vegetation in, 390-301
Gleasonian wetland, 238
Gleization, 168
Gleying, 168

Global carbon budget, 568
Global carbon cycle:
carbon budgets for peatlands, 578-579
carbon sequestration, 568-570
climate change feedbacks, 577-578
methane emissions, 570-577
peat storage and global carbon budget,
568

wetlands in, 567-579
Global Lakes and Wetlands Database
(GLWD),48

Glycolysis, 196. See Fermentation
Goals for restoration, defining, 623-624
Gravel bed, 677-678
Gray heron {Ardea cinevea)^ 533

Great Black Swamp, 63-64
Great blue heron {Ardea herodias)^ 533
Great Dismal Swamp, 8, 59-60, 114, 480
Great Kankakee marsh, 63
Great Lakes Wetlands/St. Lawrence
Lowlands, 66-68
Great Plains playas, 62-63
Green algae, 290
Greenhouse gases:

carbon dioxide, 565, 566
methane, 444, 566, 570, 572-577
nitrous oxide, 186, 567
water vapor, 566
Greenhouse gas emissions:
freshwater marshes, 365-366
from treatment wetlands, 687-688
Groundwater, 134-137
Darcy’s law, 137
recharge and discharge wetlands,
134-137

and runoff/streamflow composition,
206

water quality, 206-207
Groundwater depression wetland, 136,
137

Groundwater flows, 154-155
Groundwater slope wetland, 136, 137
Grzimek, Bernhard, 87
Guilds, 236

Gulf of Mexico, 189-192
Gulf of Mexico oil spill (2010), 66

Habitat Evaluation Procedure (HEP),
546-547

Habitat loss, mitigating, 593-597
measuring success of, 594-596
mitigation banks, 596
Hackensack Meadowlands, New Jersey,
614-616

Haliclona implexiformis^ 227
Halophytes, 323
Hammock, 57, 318

Hannah Bay Bird Sanctuary (Canada), 69
Hemists, 167

HEP (Habitat Evaluation Procedure),

546- 547

Herbivory decomposition, 361-362
Heron, 533

HGM (hydrogeomorphic) analysis,

547- 550

HGM (hydrogeomorphic) wetland
classification, 465, 467-468
High marshes, 266
Highway construction, 487-488
Histosols, 167. See also Organic soils
HLR (hydraulic loading rate), 669-670
Hong Kong Wetland Park, 99, 100
Horticulture, natural succession vs.,
633-635

Houghton Lake, Michigan, 651-653
Hudson-James Bay lowlands, 68-70
Human impact on wetlands, 4-8, 483.

See also Wetland management
Humification, 167. See also Decomposition
Hurricanes:

Andrew, 329-330

effect on mangrove swamps, 329-330
Katrina, 66, 542-543
Rita, 66

Hussein, Saddam, 598
Hydrarch succession, 228, 229
Hydraulic civilizations, 5, 479
Hydraulic conductivity, 163-165
Hydraulic loading rate (HLR), 669-670
Hydric soils, 38, 161, 513-514, 627-630
Hydrochory, 394
Hydrodynamics, 467
Hydroelectric power dams, 77
Hydrogen sulfide, 194, 195
Hydrogeomorphic (HGM) analysis,
547-550

Hydrogeomorphic (HGM) wetland
classification, 465, 467-468
Hydrogeomorphology, 112
mangrove swamps, 314-318
pannes, 267, 268
tidal creeks, 267, 268
tidal salt marshes, 265-268
Hydrologic exports, 208
Hydrologic modifications, 486-487
Hydrologic pathways, 208
Hydrology, 30, 111-155, 514
accumulation of organic material,
151-153

biotic control of, 113-115
creation/maintenance of, 625-627
cypress swamps, 377-381
evapotranspiration, 137-142
freshwater marshes, 343-344
freshwater swamps, 377-381
groundwater, 134-137
hydroperiods, 115-121
importance of, 112-115
nutrient cycling and availability, 153
and peatlands development, 416-419
precipitation, 126-127
and primary productivity, 146-151
and productivity, 146-151
recurrence interval, 133-134
red maple swamps, 381
renewal rate/turnover rate of water,

125

riparian ecosystems, 381-384
seiches, 142-144
and species richness, 145-146
study techniques for, 153-155
surface flow, 127-133
of tidal salt marshes, 266

Index 727

tides, 142, 143
in treatment wetlands design,

668-670

vegetation composition and species
richness, 145-146
water budget, 119, 121-125
and wetland fimction, 143-153
white cedar swamps, 381
Hydrologic restoration, 598-604
Florida Everglades, 598-602
Mesopotamian Marshlands, 598,

603-604

Hydroperiods, 115-121
definitions of, 115
and pulsing water levels, 119
in treatment wetlands design, 669
year-to-year fluctuations in, 119-121
Hydrophytes, 215, 216
Hydrophytic vegetation, 38
Hypoxia, 189-192

Igapo, 77

Illinois Drainage Levee Act, 482
Illinois Farm Drainage Act, 482
India, 94-95

Indian freshwater marshes, 93, 94
Indian Ocean tsunami (December 2004):
coastal restoration following, 618
mangrove swamps and, 541-542
Inflow, 128,208
Inland marshes, 266
Inland wetlands:
ecosystem, 456

effects of climate change on, 582-584
Inputs, 208
Interception, 127
Intermittently exposed, 115
Intermittently flooded, 115
Intermittent subsystem, 461, 462
International legal protections, 521-524
North American Waterfowl Management
Plan, 522-524

Ramsar Convention, 521-523
International Ramsar Convention
classification system, 465, 466
International Union for the Conservation of
Nature and Natural Resources
(lUCN), 38-39
Interstitial perennials, 236
Intertidal subsystem, 461, 462
Intrasystem cycling, 180-181, 208
Introducing vegetation:
forested wetlands, 630
freshwater marshes, 630
mangrove swamps, 633
natural succession vs. horticulture,
633-635

planting techniques, 632-633
salt marshes, 633

Invasive species:

in Florida Everglades, 600, 602-603
freshwater marshes, 353-354
tidal salt marshes, 265
Inventory, see Wetland inventory
Invertebrates:

freshwater marshes, 354-355
freshwater swamps, 395-396
Iraq, 5,89, 598,603-604
Iron:

and gleization, 168
transformation of, 192
transformations, 192
Iron bacteria, 192
Iron masses, 170
Irregularly exposed, 115
Irregularly flooded, 115
Irrigation, 62
Isolated wetlands, 28
Issyk Kul, 95

lUCN (International Union for the

Conservation of Nature and Natural
Resources), 38-39

Jacana spinosa (Northern jacana), 74
Jurisdictional wetlands, 41-42

Kahikatea Swamp, New Zealand, 91
Keeling, C. David, 565
Kennedy, Anthony, 520
Keoladeo National Park (Bharatpur, India),
93, 94

Kesterson National Wildlife Refuge,
California, 401
Kissimmee River, 58
Knees, 393-394
Konik horses, 79-81
Koontz, Coy, 520

Koontz V. St. Johns River Water Management
District., 520-521

Krefeld system (Max-Planck- Institute
process), 648
Kyrgystan, 95

Lacustrine system, 461, 462
Lacustrine wetlands, 116, 117
Lagoon, 34

Laguna de Terminos (Campeche, Mexico),
71

La^guncularia (white mangrove):
in basin mangroves, 317
salinity control, 323
Lalce Drummond, 59
Lalce-edge swamps, 378, 379
Lake Erie, 64

Lake Michigan dunes, 230-232
Landfill leachate wetlands, 668
Landscape patterns, 250-251
Lapland, 8

Larch string bog and fen, 428
Larix laricina (tamarack), 430
Legal definitions of wetlands, 39-42
Legal protection of wetlands, 503-524
Clean Water Act, 507-509, 518
Coastal Zone Management Program, 518
early Executive Orders, 504, 506
Emergency Wetlands Resource Act, 518
international, 521-524
major federal laws, directives, and
regulations, 505-506
National Academy of Science studies,
516-517

National Flood Insurance Program, 518
no net loss concept, 507, 591, 593, 595
North American Waterfowl Management
Plan, 522-524

North American Wetlands Conservation
Act, 518

Ramsar convention, 521-523
“swampbuster” provisions of 1985 Food
Security Act, 509
“takings” issue, 518
in United States, 504-521
U.S/ Supreme Court decisions, 519-521
wetland delineation guidelines, 509-516
Legal success, 594
Lenticels, 220, 325
Lignocellulose, 281
Limiting nutrients:

freshwater marshes, 364-365
peatlands, 426, 427
Limnetic subsystem, 461, 462
Limnogenous peatland, 423, 424
Linear directed change (succession), 234
Liner, 676-677
Linnaeus, Carl, 8

Literary references to wetlands, 79
Littoral subsystem, 461, 462
Littorina irrorata (snail), 284
Llanos, 7S-76
Loading rate, 669-670
Losses (hydrologic exports), 208
Loss of wetlands:

by state in United States, 701-703
worldwide, 48-50
Lotus, 83
Louisiana:

Cajuns, 5, 6

crayfish harvesting in, 10
cypress swamps, 482
energy and economic analyses of
wetlands, 554-555
Hurricane Katrina, 66, 542-543
landscape patterns in wetlands, 250
tidal salt marsh, 258
Louisiana Delta, 64-66
Low chroma, 171

Lower perennial subsystem, 461, 462

728 Index

Low marshes, 266

Lucas V. South Carolina Coastal Council^
518

Lythrum salicaria (purple loosestrife),

353

Maintenance costs of treatment wetlands,
689

Mallard {Anas platyrhynchos)^ 532
Mammals:

freshwater marshes, 357
peatlands, 433-434
tidal freshwater wetlands, 295
tidal salt marshes, 276-277
Management by objective, 491-499
for agriculture and aquaculture,
497-499

for flood control and stormwater
protection, 499

for waterfowl and wildlife, 493-496
for water quality enhancement, 499
Managing wetlands, see Wetland
management
Manchurian wild rice, 9
Mangal, 34
Manganese, 192
Manganese oxides, 168
Mangrove, 34

Mangrove swamps, 311-334

adaptations of vegetation, 322-326
consumers, 326-327
ecosystem function, 327-334
effects on estuaries, 332-334
in Ganges Delta, 94-95
geographical extent, 312-314
hurricane effects, 329-330
hvdrodynamic classification, 315-318
hydrogeomorphology, 314-318
introducing vegetation, 630
organic material storage and export,
330-332

primary productivity, 327-329
response to sea-level rise, 330
restoration of, 616-618
soil acidity, 319
soils and salinity, 319
vegetation, 319-326
zonation, 320-322
Manning equation, 132
Marceno, 497
Marine system, 460, 461
Marshes, 33, 34. See also specific types, e.g.:
Tidal marshes
Marsh Ai'abs, 5, 89
Maryland, 258
Masses, 170
Matrix perennials, 236
Matsalu State Nature Preserve, 81-82
Meander scroll, 382

Mediterranean Sea:
salt marshes in, 264
salt marsh primary productivity, 278
tidal salt marshes of, 264
Mediterranean Sea deltas, 78-79
Mekong Delta, 94, 593
Mesic riparian ecosystems, 376, 382-383
Mesopotamian Marshlands, 89, 598,
603-604

Mesotrophic peatlands, 422
Methane emissions, 198-200, 298, 299
in carbon cycle, 198-200
and climate change, 570-577
from freshwater marshes, 365
Methane oxidation, 197-198
Methanogenesis, 197, 572
Methanogens, 197-198
Methanotropic bacteria, 197
Mexico, 71-72
Miccosukee, 6, 7
Microbes, 113, 114
Middle East wetlands, 56, 89
Millennium Ecosystem Assessment, 528
Mills, Robert, 61

Mine drainage wetlands, 653-655
Mineral cycles, 241-242
Mineral extraction, alteration of wetlands in,
490-491

Mineral soils, 162, 167-171
aquic condition in, 169-170
cation exchange capacity, 166
differences between organic soils and,
163-165

hydraulic conductivity, 165
as hydric soil, 170
oxidized rhizosphere in, 168, 169
physiochemical features, 163-165
rate of formation, 168
redox concentrations in, 168, 169
reduced matrices and redox depletions,
168

Minerotrophic peatlands, 422
Mire, 34, 35
Mire ecologists, 17
Mississippi, 278

Mississippi-Ohio-Missouri (MOM) river
basin, 189

Mississippi River, 64, 664-666
Mississippi River basin, 189
Mississippi River Delta, 65
salt marshes in, 267
and sea-level rise, 580, 581
Mississippi River deltaic marshes, 260
Mitigating habitat loss, 593-597
measuring success of, 594-596
mitigation banks, 596
Mitigation ratio, 594, 595
Mitigation wetland, 593
Modifiers (for classification), 463, 464

Momoge National Nature Reserve (Jilin
Province, China), 97
MOM (Mississippi-Ohio-Missouri) river
basin, 189

Montezuma cypress ( T. distichum var.

mexicanum Gordon), 374
Moor, 33, 34

Moose River Bird Sanctuary (Canada), 69
Morphological adaptations (vascular plants),
216-220

adventitious roots, 218, 219
aerenchyma, 216-218
pneumatophores, 220
stem elongation, root adaptations, and
lenticels, 220

stem hypertrophy, 218, 219
Mosquito control at treatment wetlands,
683-686

Mottles (redox concentrations), 168-170
Movies, wetland depictions in, 9, 10
Muck, 161, 163, 167
Mucky peat, 161, 167
Mud barrens, 268

Municipal wastewater wetlands, 650-653
Munsell soil color chart, 171
Murray-Darling watershed, 593
Muskeg, 33, 34

Muslcrat {Ondatra zibethicus), 114,
276-277, 295,496
pelts from, 530

as threat to reconstructed wetland, 682
Mussels, 81
Mutualism, 226-227

Myocastor coypus (coypu, nutria), 276-277,
295, 530

NAD (nicotinamide adenine dinucleotide),
226

National Academy of Science studies,
516-517

National Flood Insurance Program, 518
National Research Council (NRC), 38
National Wedands Inventory (NWI),
472-474

National Wedands Policy Forum, 507
National Wedands Working Group
(Canada), 37
Native Americans, 5-8
Natural management, 493
Natural Resources Conservation Service
(NRCS), 40-41

Natural Resources Defense Council v.

Callaway, 40, 507
Natural succession, horticulture vs.,
633-635

Nature Conservancy, 62
Navigation, hydrologic modifications for,
487

Neap tides, 142

Index 729

Nebraska sandhills, 62-63
Nekton, 291-293
Nelumbo nucifera (water lotus), 82
Nernst equation, 174
Nerodia (water snake), 295
Netherlands, 79-80
Net marginal benefit, 556
Net primary production, 296
New England:

salt marshes, 264, 270-271
tidal salt water marshes, 264
New Orleans, Louisiana, 66, 542-543
New York City region, 614-616
New Zealand, 56, 91-92
sphagnum harvesting, 1 1
wetlands loss, 49, 50, 478, 593
Niering, Bill, 240
Nitrate transformations, 184-187
Nitrification, 183-187
Nitrobacter^ 184
Nitrogen:

and decomposition process, 282
in peatland nutrient budgets, 442-443
Nitrogenase, 187

Nitrogen cycle, 181, 183-192, 300, 301
ammonia transformations and
nitrification, 183-184
anammox, 188

dissimilatory nitrate reduction to
ammonia, 188
and hypoxia, 189-192
hypoxia and, 189-192
nitrate transformations and
denitrification, 184-187
nitrogen fixation, 187-188
nitrogen mineralization, 181
Nitrogen fixation, 187-188
Nitrogen mineralization, 181
Nitrosomonas^ 183

Nitrous oxide (N^O), 186-187, 365
Nodules, 170

Nonconsumptive use values, 545
No net loss concept, 507, 591, 593, 595
Normandy (France), 278
North American Waterfowl Management
Plan, 522-524

North American wetlands, 55-72. See also
Canada; United States
changes in, 49-55
Florida Everglades, 55-58
Great Black Swamp, 63-64
Great Dismal Swamp, 59-60
Great Kankakee marsh, 63
Great Lakes Wetlands/St. Lawrence
Lowlands, 66-68

Hudson-James Bay lowlands, 68-70
losses of, 50
Louisiana Delta, 64-66
Mexico, 71-72

Nebraska sandhills and Great Plains
playas, 62-63

Okefenokee Swamp, 58-59
Peace-Athabasca Delta, 70-71
peatlands of Canada’s central and eastern
provinces, 68
Pocosins (Carolinas), 59
prairie potholes, 61-62
river deltas, 260
San Francisco Bay, 66
swamp rivers of South Adantic coast,
60-61

wedands loss, 50

North American Wetlands Conservation
Act, 518

Northeastern China wetlands, 97
Northern Europe;

coastal marshes, mud flats, and bays,
80-82

salt marshes in, 262, 264
tidal salt marshes of, 262, 264
Northern jacana (Jacana spinosa), 74
Northern jacana [Jacana spinosa)^ 74
Northern peatlands, 165
Northern pike [Esox lucius)^ 356
NRC (National Research Council), 38
NRCS (Natural Resources Conservation
Service), 40-41
Nuphar lutea (water lily), 221
Nutria, 496

Nutria [Myocastor coypns)^ 276-277, 295,
530

Nutrients, limiting, 426, 427
Nutrient availability, 153, 164
Nutrient budgets, 208-212
components, 180
freshwater marshes, 362-364
freshwater swamps, 402, 403
peatlands, 442-445
in tidal freshwater wetlands, 301
Nutrient cycling, 153
Nutrient deficiency, 432
Nutrient limitations, 364-365
Nutrient spiraling, 405-406
NWl (National Wetlands Inventory),
472-474

Nymphaea nouchali var. caerulea (day water
lily), 85
Nyssa, 394

Nyssa aqnatica (water tupelo), 374

Obligate annuals, 236
Ob River, 93

Oceans, water quality of, 205
Ogallala Aquifer, 62
Ohio, 469-470
Oil production, 65

Okavango Delta (Bostswana, southern
Africa), 83-86

Okefenokee National Wildlife Refuge, 58
Okefenokee Swamp, 58-59
Olentaiigy River Wedand Research Park,
Ohio State University, 666-668
Oligotrophication, 423
Oligotrophic peatlands, 422
Ombrogenous peatland, 423, 424
Ombrotrophic peatlands, 422
Ondatra zibethicus (muskrat), 1 14,
276-277, 295, 496
pelts from, 530

as threat to reconstructed wetland, 682
Ontario, Canada, 55
Operating costs of treatment wetlands,
689

Opportunity costs, 551
Organic export in tidal salt marshes,
283-284
Organic material:

hydrology and accumulation of,
151-153

in mangrove swamps, 330-332
in tidal freshwater wetlands, 298-300
Organic soils, 161-162, 165-167
botanic origin of, 166-167
cation exchange capacity, 166
characteristics of, 167
classification and characteristics, 167
decomposition of, 167
differences between mineral soils and,
163-165
groups of, 167

physiochemical features, 163-165
Orinoco River delta, 74-75
Orthophosphates, 203
Outflow, 128, 208
Outwelling, 331

Outwelling hypothesis, 283-284
Overgrowth, 433
Overland flow, 128
Overwash mangrove islands, 317
Oxbow, 33, 34, 381, 382
Oxidation, 173

Oxidized rhizosphere (oxidized pore
linings), 168-170, 224
Oxygen, 171-172, 192

Paalsa peadands, 422
Pakihi, 34, 90,415

Palmer Drought Severity Index (PDSl),
344

Palm swamps, 72

Palo Verde National Park (Costa Rica),
72-74

Paludification, 416-419
Palustrine system, 461, 462
Eanicum hemitomon^ 224
Pannes, 267-268
Paiitanal, 76-77

730 Index

Panthera Tigris ti^ris (Bengal tiger), 95
Paraguay-Parana River basin, 76
Particulate organic matter (POM), 331
Pathogens at treatment wetlands, 684,

687

Patterned fens, 420-421

Patuxent River, Maryland, 298

PDSl (Palmer Drought Severity Index),

344

Peace -Athabasca Delta, 70-71
Pearsall,W. H., 228
Peat, 161-163

characteristics of, 167
and ecosystem development, 228, 230
from wetlands, 10-14
Peat accumulation, 440
Peat harvesting, 536
Peatlands, 33, 34, 90, 92, 413-445
adaptations of vegetation, 431-433
biogeochemistry, 424-427
black spruce peatlands, 430-431
bog energy flow estimates, 440-442
of Canada’s central and eastern provinces,
68

Canadian Wetland Classifleation System,
423.424

carbon budgets for, 578-579
Carolina pocosins, 431
chemistry-based classification, 422-423
classification of, 419-424, 457
and climate change, 578
consumers, 433-435
decomposition, 439
ecosystem fimetion, 436-445
and ecosystem productivity, 146
of Europe, 82-83
geographic extent, 414-416
hydrology and development of, 416-419
hydrology- based classification, 423, 424
hydroperiod of, 118-119
landscape-based classification, 419-422
nutrient budgets, 442-445
peat accumulation, 440
primary productivity, 436-438
restoration of, 605-607
soil permeability, 164
vegetation, 427-433
Peat mining, 488-489
Peat mosses, 433
Peat storage, 568
Peaty muck, 167

Pelts, animals harvested for, 529-531
Penman equation, 140-141
Perched wetland, 135, 155
Permanently flooded, 115
Permeability, 137

PET (potential evapotranspiration), 141
Philippines, 10
Phosphate mining, 490

Phosphorus, 153, 203-204, 365
Phosphorus budget, 208, 210
Phosphorus cycle, 202-204
and clay, 204

co-precipitation of phosphorus, 203-204
phosphorus release in anaerobic
conditions, 204
Photosynthesis, 195

Phragmites (reed grass), 79, 82, 95, 416,
609

and flow-through, 146
in peat composition, 164
photosynthetic efficiency, 359, 360
and reed swamps, 342
Phragmites australis (common reed), 82,
89, 609,613

as invasive species, 353-354
and reed swamps, 342
in restored wetlands, 635
Phreatophytes, 390
Physiochemical environment, 30, 112,

113

Physiognomy, 459

Physiological adaptations (vascular plants),
220-226

anaerobic respiration, 224-226
lower water uptake, 224
pressurized gas flow, 221-223
rhizosphere oxygenation, 223-224
sulfide avoidance, 224
Phytoplankton, 290
Picea mariana (black spruce), 430
Piezometers, 154-155
Pine {Pinus serotina)^ 59
Pitcher plant {Sarracenia purpurea)^ 432,
433

Plants, see Vegetation

Plant species fimctional groups (succession),
235-236
Playas, 34, 63
Pleistocene, 62

Pneumatophores, 219, 220, 325
of freshwater swamps vegetation,
393-394

in mangrove swamps, 325
Pocosins (Carolinas), 34, 59, 415-416, 431
Point bar, 382
Pokelogan, 34

Polar Bear Provincial Park (Ontario,
Canada), 68-70
Pollution, see Water pollution
POM (particulate organic matter), 331
Pond cypress {Taxodium distichum var.

imhricarium)^ 374, 375, 387
Pools, 208
Poor fens, 422
Poor swamp forest, 428
Pore linings, see Oxidized rhizosphere
Porosity, 163

Potential evapotranspiration (PET), 141
Pothole, 34

Prairie potholes, 61-62, 119, 120
Prairie Pothole Region (PPR), 582-583
Precipitation, 126-127
measuring, 154
and runoff, 206
Pressurized gas flow, 221-223
Primary productivity, 146-151
freshwater marshes, 358-361
freshwater swamps, 397-402
hydrology and, 146-151
in mangrove swamps, 327-329
peatlands, 436-438
in tidal freshwater wetlands, 295-297
tidal salt marshes, 277-281
Prop roots, 219, 220, 325
Provisioning ecosystem services, 529-538
animals harvested for pelts, 529-531
endangered and threatened species,
536-538

fish and shellfish, 534-535
peat harvesting, 536
timber and vegetation harvest, 535-536
waterfowl and other birds, 531-534
Puerto Rico, 328

Pulse stability concept, 146, 148, 233,
244-245

Pulsing water levels, 119
Puma concolor coryi (Florida panther), 58
Purple loosestrile {Lythrum salicaria)^ 353
Python molurus hivittatus (Burmese
python), 601-602

Qinghai-Tibetan Plateau, 97, 98
Qualdng bog, 417
Qualdng bog succession, 417-418
Quantifying ecosystem services, 545-559
ecological valuation, 546-550
economic evaluation, 550-558
emergy analysis, 555-556
energy analysis, 553-555
Faustian bargain in, 559
habitat evaluation procedures, 546
hydrogeomorphic analysis, 547-550
opportunity costs, 551
problems and paradoxes of, 558-559
replacement value, 551-553
valuing goods and services, 556-558
willingness-to-pay methods, 551
Quickflow, 128-129

Raised bogs, 417-420
Ramsar Convention, 38-39, 521-523
Ramsar Convention classification system,
465,466

Kapanosv. United States^ ^\9S2Q
Rapid stem elongation, 220
Rating curve, 130

Index 731

Rating wetlands, 469-471
in Florida, 470-471
in Ohio, 469-470
in Washington State, 469
Rational runoff method, 129-130
Raupo, 33
Raupo swamp, 34

RCC {river continuum concept), 404-406
Recharge wetlands, 134-137, 155
Recurrence interval, 133-134
Red-billed duck (yAnas erythrorhynchos)^ 88
Red mangrove {Rhizophora):
adaptations, 324
inundation tolerance, 315
mutualism, 227
productivity, 331
prop roots, 220
speciation, 312
viviparous seedlings, 226
Red maple {Acer rubvum)^ 59, 376
in red maple swamps, 388
root adaptations, 220
root systems, 395
swamp productivity /hydrology
relationship, 147
Red maple swamps:
geographic extent, 376
geomorphology, 381
hydrology/productivity relationship,

147

vegetation, 388-389
Redox concentrations, 168-170
Redox depletions, 168, 170
Redoximorphic features, 167-168
Redox potential, 153, 173
Reduced matrices, 168, 170
Reduction, 173

Reduction/oxidation (in soils), 171-176
Reed grass {Rhragmites)^ 33, 79, 82, 95,
416, 609

and flow-through, 146
in peat composition, 164
photosynthetic efficiency, 359, 360
and reed swamps, 342
Reedmace, 33
Reedmace swamp, 34
Reedswamp, 33, 34, 342
Reference wetland, 548-549, 594
Regularly flooded, 115
Regulation of ecosystem services,

539- 543

flood mitigation, 539-540

storm abatement and coastal protection,

540- 543

Remote platforms, 471
Remote-sensing platforms, 471-472
Renewal rate, 125, 242-244
Replacement value, 551-553
Replacement wetland, 593

Reptiles:

freshwater swamps, 396-397
peatlands, 434

in tidal freshwater wetlands, 294-295
Residence time, 125
Resource spiraling, 405-406
Respiration, 195

Restoration and creation of wetlands,
591-638

after Indian Ocean tsunami, 618
agricultural land restoration, 596-597
coastal marshes, 630
coastal restoration, 607-622
deepwater swamps, 630
defining goals, 623-624
Delaware Bay salt marsh, 608-613
deltas, 619

ecological principles for, 638-639
estimating success of, 635-638
exotic or undesirable species, 634-635
Florida Everglades, 598-603
forested wetland restoration, 598
forested wetlands, 630
freshwater marshes, 630
hydrologic and water quality restoration,
598-604

hydrology creation / maintenance ,
625-627

mangrove swamps, 616-618, 633
Mesopotamian Marshlands, 598,
603-604

mitigating habitat loss, 593-596
natural succession vs. horticulture,
633-635

New York City region, 614-616
peatland restoration, 605-607
planting techniques, 630-633
rivers, 619-620

River Skjern, Denmark, 620-622
salt marshes, 607-616, 629
site selection, 625
soils, 627-630
starting points for, 591-592
terminology related to, 592-593
Retention rates in treatment wetlands
design, 673-674
Reverse estuary, 89
Rheotrophic peatlands, 422
Rhine River delta, 79-80
Rhizohium^ 188
Rhizophora (red mangrove):
adaptations, 324

adaptations to waterlogging/flooding,
219

inundation tolerance, 315
mutualism, 227
plant zonation, 321
productivity, 331
prop roots, 220

rhizosphere oxygenation, 224
salinity control, 323
speciation, 312
viviparous seedlings, 226
Rhizophora matigle^ 227, 321
Rhizosphere oxygenation, 223-224
Rlione River Delta, 5, 6, 78, 79, 278
Ribbed mussel {Geitkensia demissa)^ 282
Rice Lalce (Crow Wing County,

Minnesota), 7
Rich fens, 422
Rich swamp forest, 428
Riparian ecosystems, 35

arid and semiarid riparian forests, 390
arid ecosystems, 377, 384
flood pulse concept, 406
geographic extent, 376-377
geomorphology and hydrology,

381-384

mesic ecosystems, 376, 382-383
river continuum concept, 404-406
and river exchanges, 404-406
southeastern U.S. bottomland forests,
389-390

vegetation, 389-390
Riparian wetlands, 132-134, 152
Rivers. See also specific rivers

restoration of wetlands, 619-622
water quality, 206-207
River continuum concept (RCC), 404-406
River diversion wetlands, 663-668
River exchanges:

flood pulse concept, 406
and riparian ecosystems, 404-406
river continuum concept, 404-406
Riverine mangroves, 317, 318
Riverine system, 461, 462
Riverine wetlands, hydroperiod of, 117-118
River Murray, New South Wales, Australia,
199

River red gum {Eucalyptus camaldulensis)^
90

River Skjern, Denmark, 620-622

River water, chemical concentrations of, 205

Rochefort, Line, 605

Rock-reed filter, 677

Roots;

adaptations of, 220
adventitious, 218, 219
of freshwater swamps vegetation, 395
morphological adaptations of, 220
prop roots, 325
shallow roots, 395
Root-zone method

{ Wurzelraumentsor^un^)^ 649-650,
677

Ruderal annuals, 236
Runoff, 127-130, 206, 207
Rtippia (wigeon grass), 268

732 Index

Sagittaria lancifolia^ 365

Sa^ittaria latifolia^ 226

St. Lawrence River, 55

St. Lawrence River Lowlands, 66-68

Salicornia^ 79

Salinity:

mangrove swamps, 319
salt marshes, 269-270
tidal salt marshes, 269-270
Salinity control, 323, 325
Salix (willow), 219
Salt, 9, 205
Salt exclusion, 323

Salt marshes, 35. See also Tidal salt marshes
conditions for development of, 259
die-off of, 284-285
inland, 349-350
introducing vegetation, 630
and Louisiana delta, 64-65
restoration of, 607-616
vegetation zonation, 272
Salt marsh cordgrass {Spartina):
and flow-through, 146
in inland marshes, 280
Salt marsh hay {Spartina patens)^ 4.97
Salt secretion, 323, 325
Sand barrens, 268
Sandpiper (Scolopacidae), 294
San Francisco Bay, 66
Saprists, 167

Sarvacenia purpurea (pitcher plant), 432,
433

Saturated, 115

Sawgrass {Cladium jamaicense)^ 57,

602

Scalia, Antonin, 520
Schoenoplectus tabernaemontani
(soft-stemmed bulrush):
in colonized wetlands, 634
nutrient limitations, 365
Scirpus americanus (American bulrush), 67,
68

Sclerophylly, 432

Sea-level rise, mangrove swamp response to,
330

Seasonally flooded, 115

Seawater, chemical characteristics of, 205

Seca, 76

Secondary treatment, 650
Sedge {Carex, Cladium)^ 147
Sedge meadow, 33, 35
Seed banks, 234-235,627
freshwater marshes, 350-351
tidal freshwater wetlands, 289-290
Seed germination and dispersal, 394-395
Seep wetland, 135
Seiches, 142-144
Seidel, Kathe, 648
Self-design, 245-249

Self-design wetlands, 591, 609, 612, 613,
618,633,634,637-638
Self-organization, 245
Semianadromous species, 292
Semiarid riparian forests, 390
Seminole, 6-8

Semipermanently flooded, 115
Services, 527
Sesarma^ 326
Shallow roots, 395
Shallow root systems, 220
Shelford, Victor, 230
Shelford curve, 198
Shellfish, 534-535
Shrub-scrub, 35
Significant nexus, 519, 520
Sinks, 181, 182

Site selection for wetland creation, 625
Skoagon Chippewa, 6
Slough, 35, 381, 382
Slow-flowing cypress strands, 378, 381
Snail {Littorina irrorata)^ 284
Snow goose, 67
Snow goose {Chen)^ 114, 682
Soft-stemmed bulrush {Schoenoplectus
tabernaemontani):
in colonized wetlands, 634
nutrient limitations, 365
Soil(s), 161-176

differences between organic and mineral
soils, 163-165
equic condition in, 169-170
hydric, 170

in mangrove swamps, 319
mineral, 162-165, 167-171
muck, 163
organic, 161-167
oxidized rhizosphere in, 168, 169
peat, 162-163
rate of formation, 168
redox concentrations in, 168, 169
reduced matrices and redox depletions,
168

reduction/oxidation in, 171-176
in restoring/creating wetland, 627-630
salt marsh, 269
tidal salt marshes, 269
for treatment wetlands, 676-679
in treatment wetlands design, 678-679
types of, 161
Soil acidity, 319
Soil chemistry, 678-679
Soil color charts, 171. See also Munsell soil
color chart
Soil profiles, 234

Solid Waste A^fency of 'Northern Cook County
(SWANCC) V. U.S. Army Corps of
En^ineers^ 519

Soligenous peatland, 423, 424

Sources:

of chemicals, 181, 182
of nutrients, 181, 182
South American wetlands, 56, 73-78
Amazon, 77-78
Llanos, 75-76
Orinoco River delta, 74-75
Palo Verde National Park, 73-74
Pantanal, 76-77
tidal marshes, 265

South Atlantic coast, swamp rivers of,
60-61

Southeastern Europe inland deltas, 82, 83
Southeastern U.S. bottomland forests,
389-390

Southern Asia river deltas, 93-95
Sparrow, 294

Spartina (salt marsh cordgrass):
and flow-through, 146
in inland marshes, 280
Spartina alterniflora:

anaerobic respiration, 225
for coastal marsh restoration, 607
in inland marshes, 280
and marsh restoration, 630
as pioneer species, 235
in restored wetlands, 635, 639
rhizosphere oxygenation, 223
root photomicrograph, 218
in salt marsh ecosystem, 270, 273
and salt marsh restoration, 613, 615
sulfiir tolerance, 224
tidal range and, 150-151
Spartina anglica^ 630
Spartina patens (salt marsh hay), 497
Spartina townsendii^ 630
Species diversity, 350-353
Species richness, 145-146
Spha^num^ 11
and acidification, 432
and bog acidity, 426
cation capacity, 165
decomposition, 439
in northern peatlands, 166
and organic soil density, 163
as peat-building plant, 427-428
and peatlands restoration, 606, 607
primary productivity, 436-438
and quaking bog succession, 417
waterlogging tolerance, 431
Spha^num-h\2rA:. spruce-beatherleaf bog
forest, 429
Sphagnum bog, 29

Spha_0num-\c2.\hc.v\c^i-Kalmia-spruce
heath., 429

Sphagnum ma^ellanicum., 437, 438
Spit, 266
Sponges, 227
Spring tides, 142

Index 733

Spring wetland, 135

SRP (soluble reactive phosphorus), 203

Standing stocks, 208

Stem elongation, 220

Stemflow, 127

Stem hypertrophy, 218-219

Storm abatement, 540-543

Stormwater protection, 499

Strand, 35

Streams, 206-207

Streamflow, 128, 131-132, 206, 207
Streamside marshes, 266
String bogs, 420-421
String fen, 421

Study techniques (hydrology), 153-155
Subclass, 463

Subsistence as cultural ecosystem service,
545

Subsurface-flow constructed wetlands, 648,
649, 679

Subsystems, 461, 462
Subtidal, 115

Subtidal subsystem, 461, 462
Succession, 227-251

allogenic vs. autogenic, 227-235
centrihigal organization concept,
238-240

community concept and continuum idea,
232-234

ecosystem development, 240-251
environmental sieve model, 236-238
Lake Michigan dunes, 230-232
landscape patterns, 250-251
linear directed change, 234
models of, 235-240
plant species hinctional groups,

235-236

seed banks, 234-235
strategy of, 242-249
Sulfate reduction, 194
Sulflde avoidance, 224
Sulfide oxidation, 194-195
Sulfide toxicity, 195
Sulfur-carbon interactions, 201-202
Sulfur cycle, 193-195
sulfate reduction, 194
sulfide oxidation, 194-195
sulfide toxicity, 195

Sundarbans (India and Bangladesh), 94-95
Supreme Court decisions, see U.S. Supreme
Court decisions
Surface flow, 127-133

channelized streamflow, 130-132
floods and riparian wetlands, 132-134
watersheds and runoff, 127-130
Surface-flow constructed wetlands, 648,
649, 679

Surface inflow, 128

Surface water depression wetland, 135, 136

Surface water slope wetland, 136-137
Sustainable cultures in wetlands, 5-8
Suwannee River, 58
Swallows (Hirundinidae), 294
Swamps, 33, 35, 289, 424. See also specific
types, e.£.: Mangrove swamps
“Swampbuster” provisions (1985 Food
Security Act), 509
Swamp gas (marsh gas), 197
Swamp Land Acts, 480-481, 504
Swamp rivers of South Atlantic coast,
60-61

SWd.NCC (Solid Waste Agency of Northern
Cook County) v. U.S. Army Corps of
Engineers, 519
Systems, 460

Taiwan urban wetland park, 99, 100
“Taldngs” issue, 518
Tamarack {Larix laricina), 430
Taxodium (cypress), 57
adaptations to waterlogging/flooding,
219

and buttresses, 394
longevity, 395

Taxodium distichum (bald cypress), 220,
374, 375, 387

Taxodium distichum var. imbricarium
(pond cypress), 374, 375, 387
Taxodium distichum var. mexicanum
Gordon (Montezuma cypress),

374

Teal, John, 281,283
Telmatology, 17
Temporarily flooded, 115
Terrace, 383
Terrestrialization, 416
Tertiary treatment, 668
Thermo-osmosis, 395
Thiobacillus, 194

Thoreau’s Bog (Massachusetts), 440, 442,
443

Thornthwaite equation, 139-141
Threatened species, 536-538
Throughfall, 127, 154
Tidal creeks, 267, 268
Tidal freshwater marsh, 35
Tidal freshwater wetlands, 285-301
amphibians and reptiles, 294-295
birds, 293-294
consumers in, 290-295
ecosystem function, 295-301
energy flow, 297-298
floating marshes, 288
mammals, 295

marsh vegetation, 285, 287-288
nekton, 291-293
new marshes, 288
nutrient budgets, 301

organic import and export, 298-300
primary productivity, 295-297
seed banks, 289-290
swamps, 289
vegetation, 285-290

Tidal marshes, 259-301. See also Tidal salt
marshes

consumers in freshwater wetlands,
290-295

ecosystem function of freshwater
wetlands, 295-301
estimated area of, 262
freshwater wetlands, 285-301
salinity, 269-270
soil, 269

vegetation, 285-290
Tidal salt marshes, 261-285
aerial habitat, 274-275
aquatic habitat, 276
Arctic, 262
Australasia, 265
benthic habitat, 276
consumers, 274-277
decomposition and consumption,
281-283

die-off of salt marshes, 284-285
eastern North America, 264
ecosystem function, 277-286
geographic extent of, 262-265
hydrogeomorphology of, 265-268
hydrology, 266
mammals, 276-277
marsh development, 266-267
Mediterranean, 264
northern Europe, 262, 264
organic export, 283-284
pannes, 267, 268
primary productivity, 277-281
salinity, 269-270
soil, 269

South America, 265
tidal creeks, 267, 268
tropics, 265
vegetation, 270-274
western North America, 264-265
Tidal subsystem, 461, 462
Tides, 142, 143
Tigris-Euphrates Basin, 89
Tigris River, 89
Timber, 535-536
Timber industry, 59-60
Todania ignis, 227
Tomatoes, 218

Topogenous development, 419
Topogenous peatlands, 423, 424
Transformers (of chemicals), 181, 182
Transitional peatlands, 422
Translocation, 208
Transportation, 487

734 Index

Treatment wetlands, 647-692
agricultural stormwater treatment
wetlands, 659-662
agricultural wastewater wetlands,
662-663

atti'acting wildlife after construction,

683

basin morphology, 670-671
capital costs, 688-689
chemical loadings, 671-676
classification of, 648-668
comparison of wetlands with

conventional technology, 690-692
defined, 593
design of, 668-682
economics and values of, 688-691
greenhouse gas emissions, 687-688
hydrology, 668-670
landfill leachate wetlands, 668
management after construction, 682-691
mine drainage wetlands, 653-655
mosquito control, 683-686
municipal wastewater wetlands, 650-653
operating and maintenance costs, 689
pathogens, 684, 687
river diversion wetlands, 663-668
soils, 676-679

urban stormwater treatment wetlands,
655-659

vegetation, 679-682
water-level management, 687
wildlife control, 682
Tropics, salt marshes of, 265
Tundra, 578

Tupelo/guni swamps, 373-374
Turlough, 35

Turnove rate of water, 125
Turnover time, 125
Typha (cattail), 73, 74
adaptations to waterlogging/flooding,
221-222

agricultural runoff and, 491
in centriftigal organization model,

239

in colonized wedands, 635
in Everglades, 601, 602
and flow-through, 146
freshwater marsh zonation, 349
and nutrient enrichment, 659-660
nutrient limitations, 365
photosynthetic efficiency, 359-360
in restored wetlands, 635
and upland soils, 627-630
and wetland primary succession, 246-249
Typha domingensis^ 73, 74, 221-222, 491,
602, 659-660

Uca (fiddler crab), 326

Undesirable species, introducing, 634-635

U.S. Army Corps of Engineers:

Louisiana Delta conservation efforts,

wetland delineation, 509-516
wetlands defined by, 39-40
wetlands definition, 41, 512
U.S. Department of Agriculture, 40, 481
U.S. Fish and Wildlife Service:

Circular 39 classification, 457-459
Habitat Evaluation Procedure, 546-547
peatland classifications, 457
wetlands defined by, 36-37
wetlands definition, 42
U.S. Fish and Wildlife Service Waterfowl
Protection Area, 62
U.S. National Academy of Science, 38
U.S. Supreme Court decisions, 519-521
U.S. wetlands, 56

CAassification of Wetlands and Deepwater
Habitats^ 460-464, 466
Clean Water Act, 507-509, 518
coastal wetlands, 260-262
Coastal Zone Management Program, 518
drainage history, 480-482
early Executive Orders, 504, 506
Emergency Wetlands Resource Act, 518
estimates of, at different times, 51
estimates of changes in, 51
Florida Everglades, 55-58
Great Black Swamp, 63-64
Great Dismal Swamp, 59-60
Great Kankakee marsh, 63
legal protections, 504-521
loss of wetlands, 51, 52
Louisiana Delta, 64-66
major federal laws, directives, and
regulations, 505-506
mangrove swamps, 313-314
National Academy of Science studies,
516-517

National Flood Insurance Program, 518
National Wetlands Inventory, 472-474
Nebraska sandhills and Great Plains
playas, 62-63

no net loss concept, 507, 591, 593, 595
North American Wetlands Conservation
Act, 518

Okefenokee Swamp, 58-59
Pocosins (Carolinas), 59
prairie potholes, 61-62
San Francisco Bay, 66
“swampbuster” provisions of 1985 Food
Security Act, 509

swamp rivers of South Atlantic coast,
60-61

“takings” issue, 518

tidal freshwater wetlands, 285

U.S/ Supreme Court decisions, 519-521

wetland delineation guidelines, 509-516

United States v. Holland^ 507
United States v. Riverside Bayview Homes,
Inc., 40, 508

Upper perennial subsystem, 461, 462
Urban stormwater treatment wetlands,

655-659

Urban wetland parks:
in China, 98-99
in Taiwan, 99, 100

Vallisneria gigantea, 199
Valuing ecosystem goods and services,
556-558
Varzea, 35, 77

Vascular plant adaptations, 216-227
adventitious roots, 218, 219
aerenchyma, 216-218
anaerobic respiration, 224-226
lower water uptake, 224
morphological, 216-220
mutualism and commensalism, 226-227
physiological, 220-226
pneumatophores, 220
pressurized gas flow, 221-223
rhizosphere oxygenation, 223-224
stem elongation, root adaptations, and
lenticels, 220

stem hypertrophy, 218, 219
sulfide avoidance, 224
whole plant strategies, 226
Vazante, 76
Vegetation, 215-227

allogenic vs. autogenic succession,
227-235

cypress swamps, 386-388
ebullition and gaseous transport in,
200-201

ecosystem development, 240-251
and evapotranspiration, 141-142
freshwater marshes, 348-354
freshwater swamps, 385-395
in glaciated regions, 390-301
harvesting as provisioning ecosystem
service, 535-536

hydrology and composition of, 145-146

introducing, 630-635

mangrove swamps, 319-326

models of succession, 235-240

morphological adaptations, 216-220

mutualism and commensalism, 226-227

peatlands, 427-433

physiological adaptations, 220-226

red maple swamps, 388-389

riparian ecosystems, 389-390

succession, 227-251

tidal freshwater wetlands, 285-290

tidal salt marshes, 270-274

treatment wetlands classification by, 650

in treatment wetlands design, 679-682

Index 735

tree adaptations, 391-395
vascular plant adaptations, 216-227
white cedar swamps, 388
Vernal pool, 35, 117
Vietnam, 593, 616-617
Viviparous seedlings, 226, 325-326
Vivipary, 226
Vleis, 35

Volatilization, 183
Vulture (Cathartidac), 294

Wadden Sea, 81
Wad/wadden, 35
Washington State, 469
Wastewater treatment wetland
classifications:

agricultural stormwater treatment
wetlands, 659-662
agricultural wastewater wetlands,
662-663

general approaches to, 648-650
landfill leachate wetlands, 668
mine drainage wetlands, 653-655
municipal wastewater wetlands, 650-653
river diversion wetlands, 663-668
urban stormwater treatment wetlands,
655-659

by vegetation, 650
Water budget, 119-125, 154, 380
examples of, 122-125
major components of, 122
Water exti'action, 490-491
Waterfowl:

as provisioning ecosystem service,
531-534

wetland management by objective for,
493-496

Water-level management, 687
Water lily {Nuphar liitea)^ 221
Waterlogging, 431
Water lotus {Ndumbo nucifera)^ 82
Water pollution:

alteration of wetlands by, 491, 599,

602

and mine drainage wetlands, 653, 655
and wastewater treatment wetlands, 655,
659, 662-663, 667, 673, 691, 692
Water quality, 204-207

and climate regulation, 544-545
oceans and estuaries, 205
streams, rivers, and groundwater,
206-207

wetland management by objective for
enhancing, 499
wetlands created to improve,
see Treatment wetlands
Water quality restoration, 598-604
Florida Everglades, 598-603
Mesopotamian Marshlands, 603-604

Watersheds, 127-130
Water snake {Nerodia), 295
Water tupelo {Nyssa aquatica), 374
Water uptake, adapting, 224
Weirs, 131

West Africa mangrove swamps, 87-89
West Bengal, 94

Western Australia wedands, 90-91
Western North America:
salt marshes, 264-265
tidal salt marshes, 264-265
Western Siberian lowlands, 93
Wetlands, 3-22. See also U.S. wetlands
of Africa, 56, 83-89
of Asia, 56,93-100
of Australia, 56, 89-91
of Central America, 56, 72-74
common terms used to describe,

33-36

conservation of, 16-17
defining, 28-32

distinguishing features of, 28-31
and ecotourism, 13-15
of Europe, 56, 78-83
food from, 9-10
formal definitions of, 36
global extent of, 45-48
historical human impact on, 4-7
legal definitions of, 39-42
literary references to, 79
management of, 21-22
of Middle East, 56, 89
of New Zealand, 56, 91-92
of North America, 55-72. See North
American wetlands
peat and building materials from,

10-14

regional, 55, 56
of South America, 56, 73-78
specialization in study of, 17-21
sustainable cultures in, 5-8
as term, 27-28

web pages related to, 705-707
worldwide losses of, 48-50
Wetland classification, 455-474

Canadian Wetland classification System,
465,466

Circular 39 classification, 457-459
coastal wedand classification, 459-460
hydrogeomorphic wetland classification,
465,467-468

international Ramsar Convention
classification system, 465, 466
peatland classifications, 457
for rating wetlands, 469-471
reasons for, 456-457
U.S. Classification of Wetlands and
Deepwater Habitats^ 460-464, 466
for wetland inventory, 471-474

Wetland creation, 592
Wetland delineation guidelines (U.S.),
509-516

Wetland ecologists, 17
Wetland ecology, 17
Wetland enhancement, 592
Wetlanders, 5
Wetland function;

accumulation of organic material,
151-153

hydrology and, 143-153
nutrient cycling and availability, 153
and primary productivity, 146-151
vegetation composition and species
richness, 145-146
Wetland inventory, 455, 471-474
remote-sensing imagery, 472
remote-sensing platform, 471-472
U.S. National Wetlands Inventory,
472-474

Wetland management, 21-22, 477-499
for agriculture and aquaculture, 497-499
alteration of wetlands, 482-492
drainage, dredging, and filling, 482-486
early history of, 478-480
for flood control and stormwater
protection, 499

and highway construction, 487-488
hydrologic modifications, 486-487
mineral and water extraction, 490-491
by objective, 491-499
peat mining, 488-489
U.S. wetland drainage history, 480-482
for waterfowl and wildlife, 493-496
and water pollution, 491
for water quality enhancement, 499
Wetland managers, 21-22
Wetland restoration, 592. See also

Restoration and creation of wetlands
Wetland science, 17-21
Wetland scientists, 17
Wetlands of international importance, 17
Wetlands Reserve Program, 597
Wet meadow, 33, 35
Wet prairie, 33, 35

Whangamarino Wetland {North Island,

New Zealand), 90

White cedar {Chamaecyparis thyoldes)^

375

White cedar swamps:

geographic extent, 375-376
geomorphology, 381
vegetation, 388

White-faced heron {Ardea novaehollandiae),
533

White mangrove (La^uncularia):
in basin mangroves, 317
salinity control, 323
White-necked heron {Ardea cocol)^ 533

736 Index

White pine, see Kahikatea
Wigeon grass {Ruppia)^ 268
Wildlife:

at treatment wetlands, 682-683
wetland management by objective for,
493-496
Wild rice, 7

Wild rice {Zizania)^ 68, 166, 497
Willingness-to-pay, 551
Willow {Salix)^ 219
Wilma H. Schicrmcicr Olentangy River
Wetland Research Park (Ohio State
University), 246

Wilson, L. R., 228

Wind tide, 143. See also Seiches

Wisconsin, 6

Wren (Troglodytidae), 294
Wurzelraumentsorgung (root-zone
method), 649-650, 677

Xixi National Wetland Park (Hangzhou,
China), 98, 99

Yangtze Delta, China, 95
Yangtze River wetlands, 95, 97

Yellow-billed duck {Anas undulata)^ 88
Yellow-billed stork {Ibis ibis)^ 87, 88

Zabel V. Tabb, 40

Zizania (wild rice), 7, 9, 68, 166, 497

Zizania aquatica^ 7, 497

Zizania latifolia^ 9

Zizania palustris^ 68

Zonation:

in freshwater marshes, 348-349
in mangrove swamps, 320-322
Zooplankton, 298

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